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THE EFFECTS OF TEMPERATURE, SUSPENDED SOLIDS, AND ORGANIC CARBON ON COPPER TOXICITY TO TWO AQUATIC INVERTEBRATES CHAD J. BOECKMAN∗ and JOSEPH R. BIDWELL Oklahoma State University, Department of Zoology, Ecotoxicology and Water Quality Research Laboratory, Stillwater, OK 74078 (∗ author for correspondence, e-mail: [email protected]; Tel: (405) 744-9691, Fax: (405) 744-7824)

(Received 9 September 2005; accepted 9 November 2005)

Abstract. Ephemeral wetlands are characterized by temporal changes in abiotic characteristics that could ameliorate or exacerbate contaminant effects on resident species. The goal of this study was to determine the effects of temperature and naturally occurring suspended solids and organic carbon on the response of Daphnia pulex, and the calanoid copepod, Diaptomus clavipes, to a copper reference toxicant. Organisms were exposed to copper at 10, 20 and 30 ◦ C in 48-h static renewal tests with a diluent of either reconstituted laboratory water or water from a wetland that had elevated levels of both suspended solids and organic carbon. 48-h LC50 values were calculated for both total and free ion copper concentrations. When wetland water was used as the diluent, LC50 values based on total copper concentrations were significantly greater than free ion LC50 s for both species. This difference was not as great in laboratory water, indicating that binding of the metal was greater in the wetland diluent and the free ion was largely responsible for toxicity. While D. clavipes was significantly less sensitive to the metal than D. pulex (48-h LC50 for total copper in laboratory water at 20 ◦ C 607.4 μg/L vs. 10.7 μg/L, respectively), the copepod exhibited a much greater response to increasing temperature. When the Biotic Ligand Model was used to generate free ion concentrations, it was found that measured concentrations exceeded the predicted values at each test condition; however measured LC50 values for D. pulex were within a factor of two of the predicted LC50 ’s at all temperatures and in both diluents. Keywords: biotic ligand model, copper, free ion, organic carbon, suspended solids

1. Introduction The response of an organism to contaminant exposure represents an integration of both biological and environmental factors which may ameliorate or exacerbate toxicity. For example, variables such as temperature can significantly increase sensitivity to chemical stressors as demonstrated by Persoone et al. (1989), who found a 12-fold increase in the number of Daphnia magna immobilized during exposure to potassium dichromate when the temperature was raised from 7 to 28 ◦ C. Similarly, in studies with the calanoid copepod, Diaptomus clavipes, Cooney et al. (1983) demonstrated that temperature and nutritional state influenced sensitivity to the aromatic hydrocarbon, acridine. Water, Air, and Soil Pollution (2006) 171: 185–202 DOI: 10.1007/s11270-005-9036-3

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C.J. BOECKMAN AND J.R. BIDWELL

Other environmental variables, such as suspended solids, may bind certain contaminants, reducing their bioavailability and effect on aquatic organisms (McCarthy, 1983; Muir et al., 1983; Schrap and Opperhuizen, 1990; Polonsky and Clements, 1999). Van Veen et al. (2002) found that when water high in suspended solids was used as the diluent in bioassays with D. magna, tolerance to copper was more than four times greater than when a diluent free of suspended solids was used. Dissolved organic carbon may also bind some contaminants and reduce overall bioavailability (Stackhouse and Benson, 1988; Stackhouse and Benson, 1989; Meador, 1990; Wood and Shelley, 1999). Turbidity and suspended solids may also act as an exacerbating stressor. For example, because most species of cladocerans are nonselective filter feeders (Kirk and Gilbert, 1990), suspended solids may be ingested, reducing caloric intake and/or increasing body weight so the organism must expend more energy to remain at the desired depth within the water column (Zurek, 1983; Herbrandson et al., 2003a,b). Ephemeral wetlands may serve as model systems for studies of the effects of environmental variables on contaminant bioavailability for several reasons. These systems often act as collection pools for surface runoff, and as such, have the potential to receive a wide range of contaminants associated with surrounding land use (Catallo, 1993; Clark et al., 1993; Donald et al., 2005). Many such wetlands also have levels of suspended solids and organic carbon that can be quite high and fluctuate seasonally (Eriksen, 1966; Black, 1976; Colburn, 2004). Other physical-chemical parameters such as temperature can also vary both temporally and spatially, further leading to heterogeneous exposure conditions. While some of the invertebrates commonly found in these systems (e.g. daphnids) have been the subject of toxicological studies for some time, these species are normally tested under a standardized set of abiotic conditions. The biological result of the combined natural environmental stressors and influences on contaminant bioavailability remains poorly characterized. Given the importance of ephemeral wetland habitats for invertebrates, amphibians, reptiles, shorebirds and waterfowl (Hartland-Rowe, 1966; Haukos and Smith, 1994; Battle and Golladay, 2001; Dodson and Lillie, 2001; Russell et al., 2002; Wood, et al., 2003), an understanding of the risk chemical stressors could pose to these systems could be important for management purposes. In order to begin evaluating this issue, we sought to characterize the effects of temperature and naturally occurring suspended solids and organic carbon on copper toxicity to the cladoceran, Daphnia pulex, and the copepod, Diaptomus clavipes. We also sought to compare our results with those predicted from the Biotic Ligand Model (BLM), which has been used to predict metal speciation and binding based on water quality characteristics of the diluent (Di Torro et al., 2001; Santore et al., 2001).

THE EFFECTS OF ABIOTIC VARIABLES ON COPPER TOXICITY

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2. Methods 2.1. STUDY

SITE

The ephemeral wetland used in this study is located in a rural area approximately 15 km southwest of Stillwater, in Payne County, Oklahoma. The upland area is dominated by a mix of grasses including old world bluestem (Bothriochloa ischaemum) and Bermuda grass (Cynodon dactylon) (Amy Ganguli, Oklahoma State University, personal communication), and cattle occasionally graze the field surrounding the site in late spring and early summer and also sometimes use it for drinking water. The system typically fills after a 7–10 cm rain, and water will persist for several months depending on soil saturation, air temperatures, and relative humidity. When full, the dimensions of the wetland approach 0.16 ha with a maximum depth of 60–80 cm and an average depth of 30 cm. 2.2. STUDY

ORGANISMS

The calanoid copepod, Diaptomus clavipes, and the daphnid, Daphnia pulex, were used as the test species. On the day of each test, D. clavipes were collected from the wetland with a 153 μ M mesh zooplankton net attached to a metal pole. Multiple sweeps at various depths were taken until enough organisms (approximately 500) were collected for the bioassays. The animals were then transported back to the laboratory in a polyethylene container filled with wetland water, where they were directly transferred into soft reconstituted water (USEPA, 2002a) that approximated wetland water hardness, and alkalinity. Identity of the organisms was verified following keys by Pennak (1953) and Edmondson (1959). D. pulex were obtained from an in-house culture maintained at Oklahoma State University. Mass cultures were held at 23 ± 2 ◦ C and fed 1 μL of a pre-digested trout chow/yeast mixture (USEPA, 2002a) and 1 μL of Selenastrum capricornutum daily (USEPA, 2002a). Culture water was exchanged every other day, and D. pulex were generally 4 d old at time of collection. 2.3. DILUENTS The acute toxicity of copper on D. clavipes and D. pulex was assessed in two treatments, one using water from the wetland and another using soft reconstituted laboratory water described previously. The morning of the bioassay, wetland water was collected in an acid washed 8-L polyethylene container. During times of high suspended solid concentrations, the soft reconstituted water was used to dilute the wetland water, generally by 50%, which allowed the test organisms to remain visible.

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2.4. BIOASSAY

METHODS

All bioassays followed methods outlined in USEPA (2002b). Copper sulfate (CuSO4 ) (Sigma Chemical CO., St. Louis, MO) was used as a standard toxicant in side-by-side 48-h static-renewal acute toxicity tests with both diluents. Test solutions were prepared in acid washed, 500 mL glass volumetric flasks by adding a small volume (1–10 mL) of a concentrated copper solution to the diluent, followed by stirring for 1 hour. This method was employed to reduce dilution of suspended solids and organic carbon in the wetland water diluent. Test organisms were placed R 4 per 30-mL polyethylene Solo cup (Solo Cup Company, Urbana, IL), with 5 cups per concentration and 20 mL of solution per cup. All organisms were acclimated in cups containing diluent only for 24 h prior to the start of the bioassays. During this acclimation phase, organisms were fed 1 μL of an algal suspension (Selenastrum capricornutum, 1.8 × 107 cells/mL), but were not fed during the bioassays. After the 24-h acclimation phase, the organisms were transferred to a copper spiked diluent of either the soft reconstituted water or wetland water. Bioassays were conducted in environmental chambers (Percival Scientific Inc. Boone, IA) at temperatures of 10, 20, or 30 ◦ C, and three replicate tests with each organism were conducted at each temperature. These temperatures were chosen to correspond to the seasonal wetland conditions at the time the copepods were collected (10 ◦ C tests for early spring, 20 ◦ C during early summer and, 30 ◦ C during late summer), minimizing temperature acclimation periods for the D. clavipes. D. pulex cultures were given 2 days to acclimate to the bioassay temperatures. Tests were conducted under an ambient light-dark regime, approximating natural light conditions at the wetland. The nominal copper concentrations used in the tests varied based on the species tested, the diluent used, and the exposure temperature. Test solutions were prepared and exchanged daily. Numbers of surviving organisms were recorded at 24 h and 48 h, and any dead individuals were removed from the cups. 2.5. WATER

CHEMISTRY

Prior to collecting water, dissolved oxygen, pH, conductivity, and temperature were measured in the wetland with a Hydrolab Quanta probe (Hydrolab Corporation, Austin, TX.), and alkalinity and hardness were determined in the laboratory (APHA, 1998). Samples (approximately 100 ml) of both laboratory and wetland water were also collected in 125 mL polyethylene containers and preserved by acidification with phosphoric acid to a pH less than 2 for total organic carbon analysis (TOC). TOC samples were analyzed by the University of Georgia (Athens, GA) Chemical Analysis Laboratory. Total suspended solids (TSS) were determined by placing 10 ml undiluted wetland water into a pre-dried and pre-weighed aluminum pan (10 replicates/sample), heating to 103 ◦ C (±2 ◦ C) for 24 h to evaporate off the water in a drying oven, and then reweighing the residue remaining on the pans after they had cooled (APHA, 1998). TSS and TOC levels in the wetland diluent bioassays were


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calculated by dividing the known concentration of each parameter on the undiluted wetland water by the test dilution factor. For determination of copper concentrations in the test solutions, approximately 100 mL of each treatment was placed in a 125-mL polyethylene bottle and preserved by acidification with metal grade nitric acid to a pH between 1.5–2.0 for later analysis using either flame or graphite furnace atomic absorption spectroscopy on a Perkin Elmer AAnalyst 700 atomic absorption spectrometer (Perkin Elmer Inc., Wellesley, MA). Calibration standards were prepared by serial dilution of a copper standard (Ultra Scientific, Kingstown, RI) and check standards were run with each set of analyses. Free copper ion concentrations were also determined within 1 h of making each test dilution with an Accumet cupric combination probe and mV/ ion meter (Fisher Scientific, Hanover Park, Il.) following the manufacturer’s protocol for analysis. 2.6. STATISTICAL

ANALYSIS

After completion of each bioassay, two LC50 ’s (lethal concentration to 50% of the exposed organisms) were generated using the trimmed Spearman-Karber method with software available from the USEPA (www.epa.gov/EERD/stat2.htm). One LC50 was based on measured total copper concentrations determined from atomic adsorption spectroscopy, and the other was generated using the free-ion concentrations from each dilution. Initial comparison of LC50 values was completed by visual inspection of the degree to which the 95% confidence intervals overlapped. In addition, paired t-test comparisons of LC50 values between and among temperatures with a Holm-Sidak correction were performed with SigmaStat version 3.1 (Systat Software Inc., Richmond CA. USA). Regression analyses were used to determine the influence of TSS and TOC on LC50 values using SigmaPlot version 7.0 (Systat Software Inc., Richmond CA. USA). Further evaluations of acute effects levels and the influence of TOC, TSS and temperature on these effects levels were accomplished through use of the Biotic Ligand Model (BLM), (www.hydroqual.com/wr blm.html). The BLM was also used to compare measured copper free-ion values with the model-predicted values.

3. Results 3.1. DILUENT

CHARACTERISTICS

The soft laboratory water used as the diluent in bioassays ranged in pH from 7.2 to 7.6 standard units. Hardness ranged from 40 to 48 mg/L CaCO3 with an alkalinity between 30 and 35 mg/L CaCO3 . Conductivity varied between 0.105 and 0.120 mS/cm. Physical-chemical characteristics for the wetland water are presented in Table I. Water was collected for use in testing from March through October

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TABLE I Range of physical-chemical characteristics for the wetland water used as diluent for the bioassays

Species

Date collected

Temp. (◦ C)

Alkalinity (mg/L CaCO3 )

Hardness (mg/L CaCO3 )

TSS (mg/L)

TOC (mg/L)

D. pulex D. pulex D. pulex D. clavipes D. clavipes D. clavipes

6/22/04 7/28/04 8/30/04 3/11/04 4/20/04 10/25/04

10 20 30 10 20 30

25–35 32–38 32–38 10–20 22–30 33–40

32–44 38–48 43–52 10–20 30–38 44–50

73.5–190 141.6–148.5 160.3 162.3 108.8–188.0 100.0–118.0

10.4–16.4 14.0–16.9 15.79 7.78 7.48–7.56 26.2–30.1

2004. For all bioassays with D. pulex, alkalinity ranged from 25 to 38 mg/L CaCO3 with a hardness range from 32 to 52 mg/L CaCO3 . TSS varied between 73.5 and 190.0 mg/L with a TOC range from 10.4 to 16.9 mg/L. In bioassays with D. clavipes, alkalinity varied between 10 and 40 mg/L CaCO3 with hardness ranging from 10 to 50 mg/L CaCO3 . TSS and TOC ranged from 100.0 to 188.0 mg/L and 7.48 to 30.1 mg/L, respectively. 3.2. BIOASSAYS

WITH

D.

PULEX

Average 48-h LC50 values generated from tests with D. pulex at 10, 20, and 30 ◦ C, are presented in Figure 1a. Regardless of temperature, values based on total copper concentrations in the wetland water were significantly higher than those from bioassays in which soft laboratory water was the diluent. Mean total LC50 ’s ranged between 208.8 μg/L at 10 ◦ C to 186.1 μg/L at 30 ◦ C for the wetland water, while those from tests with laboratory water ranged from 9.1 to 7.2 μg/L at 10 and 30 ◦ C, respectively. When free copper ion concentrations were used to generate LC50 values, the difference between the two diluent types was less pronounced, although values from the wetland water tests were still significantly higher than the numbers from laboratory water tests at 10 and 20 ◦ C ( p = 0.002). No difference in the free ion LC50 values was observed between the two diluents at 30 ◦ C. While no significant temperature effects were apparent in the 48-h LC50 values based on either total or free copper levels for this species, there was a slight increase in the average value for total metal at 20 ◦ C for both the wetland and laboratory water. A similar trend was not observed for the free ion values. The degree of difference between total and free ion LC50 values depended on the diluent used. In wetland water, the total LC50 values were always significantly greater than the free ion values regardless of temperature ( p = 0.001, Figure 1a). For the soft laboratory water, there was no significant difference between the total


191

Figure 1. Average (n = 3 bioassays) 48-h acute LC50 values for Daphnia pulex (a) and Diaptomus clavipes (b) exposed to copper at 10, 20, and 30 ◦ C in both soft laboratory and wetland water, and BLM predicted LC50 values for D. pulex in both diluents (a). Bars represent values derived from either total measured copper concentrations or free copper ion. Error bars indicate 95% confidence intervals.

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and free ion LC50 values at 10 and 30 ◦ C, but at 20 ◦ C the free ion value was significantly lower ( p = 0.002). As observed for the total LC50 values, those based on free copper ion concentrations did not indicate a significant influence of temperature on the response of D. pulex, although at least for the wetland water, there was a trend of decreasing free ion LC50 as temperature increased (Figure 1a). 3.3. BIOASSAYS

WITH

D.

CLAVIPES

In most cases, the total and free copper LC50 values generated for D. clavipes were significantly higher than those for D. pulex, often by an order of magnitude or more (Figure 1b). As with D. pulex, LC50 values for total copper derived from tests with the wetland water were significantly greater than those from tests in the soft laboratory water. For this species, mean total LC50 ’s ranged between 1400 μg/L at 10 ◦ C to 467.5 μg/L at 30 ◦ C for the wetland water, while those from tests with laboratory water ranged from 510 μg/L at 10 ◦ C to 53.3 μg/L at 30 ◦ C. There was no significant difference between the diluents based on comparison of LC50 values derived from the free ion concentrations, although the total LC50 values from bioassays with both diluents were significantly greater than the corresponding free ion values regardless of temperature ( p < 0.01). In contrast to what was observed for D. pulex, there was a marked influence of temperature on the response of the copepods, with the 48-h LC50 values generated at 30 ◦ C significantly lower than those at the other two temperatures regardless of diluent or chemical nature of the metal (Figure 1b, p < 0.001). It is also interesting to note that the increased sensitivity of D. clavipes at 30 ◦ C occurred even though the organic carbon levels in these tests were four times greater than those at 10 and 20 ◦ C (Table I). For the wetland diluent, the average LC50 value from the copepod bioassays at 20 ◦ C was also significantly greater than that for the tests at 10 ◦ C for both total and free copper. 3.4. INFLUENCE OF TOTAL SUSPENDED SOLIDS AND TOTAL ORGANIC CARBON ON COPPER TOXICITY

Simple linear regressions were generated to evaluate the influence of TSS and TOC on the 48-h LC50 values for D. pulex (Figure 2a–d). Considering increased temperature significantly reduced median lethal concentration for D. clavipes, regressions were not produced for this species since data from all three temperatures were combined to conduct the analysis. For D. pulex, both TSS and TOC levels had a significant influence on the LC50 s for total copper concentrations (respectively, r 2 = 0.84, p < 0.0001; and r 2 = 0.91, p < 0.0001). When these environmental variables were regressed against free copper ion LC50 values, TSS (Figure 2b) exhibited a significant, yet weak relationship (r 2 = 0.29, p = 0.02), whereas TOC (Figure 2d) did not ( p = 0.055).

193


TSS vs. Tot LC50

350

30

48-h Total LC50 (u g Cu/L)

250 200 150 100 50 r 2= 0.84 y = 1.31x + 21.66

0

48-h Free ion LC50 (ug Cu/L)

a

300

TSS vs. Free ion LC50

b

25

20

15

10

5 r 2= 0.29 y = 0.04x + 7.09

0 0

50

100

150

200

0

50

TSS (mg/L)

100

150

200

TSS (mg/L) 10 oC 20 oC 30 oC

TOC vs. Tot LC50

350

30

48-h Total LC50 (ug Cu/L)

250 200 150 100 50 r 2 = 0.91 y = 14.02x - 1.39

0

48-h Free ion LC50 (u g Cu/L)

c

300

TOC vs. Free ion LC50

d

25

20

15

10

5

2

r = 0.21 y = 0.38x + 7.04 0

0

2

4

6

8

10 12 14

TOC (mg/L)

16 18

0

2

4

6

8

10 12 14

16 18

TOC (mg/L)

Figure 2. Linear regressions of total suspended solids (a and b) and total organic carbon (c and d) versus 48-h LC50 values derived from either total copper (a and c) or free copper ion (b and d) for the water flea, Daphnia pulex. Data from all bioassays across each of the three temperatures was used to generate the regression.

3.5. BIOTIC

LIGAND MODEL

The Biotic Ligand Model (BLM) was used to predict free copper ion concentrations for each species in each diluent. Measured free copper concentrations for bioassays

194


D. pulex in soft water

100

a

10

BLM predicted Cu++ (u g/L)

BLM predicted Cu++ (u g/L)

100

1

0.1

0.01

0.001

D. pulex in pond water

b

10

1

0.1

0.01

0.001

0.0001 0.1

1

10

0.0001 0.1

100

1

Measured Cu++ (u g/L)

10

100

Measured Cu++ (u g/L) 10 oC o 20 C 30 oC

100000

100000

c

BLM predicted Cu++ (ug/L)

10000

BLM predicted Cu++ (ug/L)

D. clavipes in pond water

D. clavipes in soft water

1000 100 10 1 0.1

d

10000 1000 100 10 1 0.1 0.01

0.01 0.001 0.1

1

10

100

1000


10000

0.001 0.1

1

10

100

1000

10000


Figure 3. Measured free copper vs. BLM predicted free copper concentrations for bioassays conducted with Daphnia pulex (a and b) and Diaptomus clavipes (c and d) in soft laboratory water (a and c) and wetland water (b and d). Data from all three temperatures tested were included in the plots. Dashed lines represent unity between predicted and measured copper free-ion concentrations.

with D. pulex in both soft laboratory and wetland water were significantly greater than the BLM predicted concentrations (Figure 3a and b), as indicated by the departure of these values from unity. Bioassays with D. clavipes (Figure 3c and d) exhibited better agreement with predicted free copper concentrations, however measured values for tests conducted at 30 ◦ C were still well off the line of unity. The BLM was also used to predict total LC50 values for D. pulex in both diluents and at the three different test temperatures (Figure 1a). Predictions were not made


195

for D. clavipes since it is not one of the species included in the database for the model. The BLM-predicted total LC50 values for D. pulex exposed in wetland water were significantly greater than the predicted values for the soft laboratory water at all three temperatures. In contrast to the predicted free-ion data, BLM predicted LC50 ’s for total copper were similar to the total measured values, particularly at 20 ◦ C.

4. Discussion 4.1. TOTAL

VS . FREE - ION

LC50 ’S

In all bioassays conducted with D. pulex or D. clavipes, total copper LC50 ’s in the soft laboratory water were always less than those in the wetland water. Total copper LC50 values from bioassays with D. pulex were also positively correlated with levels of TSS and TOC. These results are not surprising since organic carbon, suspended solids (especially clay particulates), carbonates, hydroxides, sulfides, and chlorides will all complex with copper and potentially decrease toxicity (Kim et al., 1999; Polonsky and Clements, 1999; Wood and Shelley, 1999; Ma et al., 2002; Sparks, 2003; Hyne et al., 2005). The relationship between free-ion (Cu2+ ) LC50 values and TSS or TOC was not significant or weak, respectively. By definition, the free-ion represents that portion of the metal which is most bioavailable (Campbell, 1995), and so capable of entering the organism and reaching an active site. The concentration of bioavailable copper (as indicated by the free-ion) needed to cause toxicity should therefore stay the same even though the total amount of metal associated with the response may change as levels of complexing agents increase. This explains why the free-ion LC50 values from bioassays in wetland water were always significantly lower than LC50 ’s based on total metal levels, regardless of the species tested. In studies investigating the effects of various water chemistry variables on copper toxicity to fathead minnows (Pimephales promelas), Welsh et al. (1993) and Erickson et al. (1996) similarly found no relationship between measured free-ion concentrations and organic carbon or suspended solid content. While the LC50 values based on total copper and free ion in the soft laboratory water were more similar than observed for the wetland diluent, the free ion values were still usually lower. This was particularly apparent in tests with D. clavipes in which free ion LC50 values were always significantly lower than total values, regardless of the diluent. Since the laboratory water still contained a suite of potential complexing agents (carbonate, hydroxide, chloride), these results are not entirely surprising. It is also possible that there was some binding of metal to the sides of the test chambers with associated loss of free ion. The reason for the apparently greater disparity in free ion and total LC50 values from bioassays with D. clavipes is not entirely apparent. Since the copepods were derived from the wetland, it is possible that some low levels of solids and/or organic carbon were introduced into the test

196


chambers as the organisms were being pipeted in. However, since the copepods were initially acclimated to the laboratory water, any residual material adhering to their body should have been washed off. The copepods were also significantly more active in the test chambers than the water fleas, and this may have promoted more interactions between metal ions and potential complexing agents. Our copper tests solutions were prepared fresh daily with at least 1 h mixing before use. In more dilute solutions such as the laboratory water diluent, equilibration between copper ions and complexing agents may need longer to occur. Kim et al. (1999) found that copper solutions mixed for 24 h before use in bioassays with Ceriodaphnia dubia were less toxic than solutions which had shorter mixing times before exposure of the organisms. They attributed this effect to a greater degree of equilibration and binding of the metal with dissolved organic matter in the treatments that had been mixed longer. 4.2. COMPARATIVE

SENSITIVITY OF THE TEST ORGANISMS

Overall, D. clavipes was much less sensitive to copper in both diluents than was D. pulex. Similar studies evaluating the comparative toxicity of a metal to these two species appear lacking. However, in separate bioassays with the organic acridine that were conducted at comparable temperatures, Southworth et al. (1978) reported a 24-h LC50 of 2.92 mg/L for D. pulex, while Cooney et al. (1983) reports a 48-h LC50 of 4.8 mg/L for D. clavipes. Thus, even with the longer exposure, D. clavipes was less sensitive to this particular chemical. It is possible that D. clavipes is generally more tolerant to chemical stressors than the daphnids due to factors such as rates of uptake and/or excretion. The differences may also be related to the source of the organisms, with the copepods collected from the field more robust than the D. pulex which were derived from laboratory cultures and not subjected to the type of environmental selective pressures that may be influencing populations of the former. 4.3. TEMPERATURE

EFFECTS

For D. pulex, temperature had little apparent influence on response of the organisms to copper, while for D. clavipes, it did. The results for the water flea contrast with those of Lewis and Horning (1991) and Heugens et al. (2003) in which Daphnia magna and D. pulex were more sensitive to metals at 26 ◦ C than they were at 20 ◦ C. Heugens et al. (2003) suggested that increased sensitivity resulted from a combination of greater metal uptake and thermal stress caused by the higher temperatures. This apparent lack of temperature influence on D. pulex may be a statistical artifact since, at least for the total copper LC50 values, there is a decrease between 20 and 30 ◦ C, indicating a trend toward increased sensitivity to the metal at the higher temperature. If more replicate bioassays had been conducted, it is possible that these differences would have been statistically significant. It is also interesting to note


197

that sensitivity of D. pulex was lowest at 20 ◦ C, indicating potential temperature stress at both 10 and 30 ◦ C. Most organisms exhibit some temperature range that is optimal for their physiological function (Wilmer et al., 2005). Since the daphnids used in these bioassays were derived from cultures maintained at 21–25 ◦ C, it is reasonable to conclude that the 20 ◦ C tests were conducted near this optimum temperature range In contrast to D. pulex, raising the bioassay temperature to 30 ◦ C resulted in a significant increase in the sensitivity of D. clavipes to copper. Cooney et al. (1983), and Cooney and Gehrs (1985) suggest that 28 ◦ C is the upper temperature threshold for D. clavipes populations in the wild, so thermal stress was probably a significant factor in these results. The copepods were least sensitive to copper at 20 ◦ C which was also suggested by the data for D. pulex. In studies investigating the effects of acridine on D. clavipes, Cooney et al. (1983), and Cooney and Gehrs, (1985) also observed a decrease in sensitivity when temperatures were raised from 16 to 20 ◦ C, followed by an increase in sensitivity when temperatures were again raised to 26 ◦ C. As discussed for D. pulex, it is possible that the organisms have a temperature optimum that is either near, or includes, 20 ◦ C, although the upper limit may be less than that for the daphnids which could account for the disparity in sensitivity to copper at the highest test temperature. The existence of such temperature optima has important implications for invertebrates that occur in ephemeral wetlands since temperatures in these systems can fluctuate significantly on both a seasonal and daily basis (Black, 1976; Serrano and Toja, 1995). As such, the degree of risk that chemical stressors pose to these systems may also exhibit similar temporal variability. For both D. pulex and D. clavipes, temperature-induced effects on sensitivity to copper appeared greatest in the wetland water as compared to the laboratory diluent. This could indicate an added stress on the organisms that is imposed by the elevated suspended solids in the wetland water which increases overall sensitivity to copper when combined with temperatures that are outside of optimal ranges. In a study investigating copper toxicity to Ceriodaphnia dubia, Ma et al. (2002) demonstrated that suspended solids could have additional adverse effects by way of physical forces or ingestion. Similarly, Zurek (1983), Kirk and Gilbert (1990), and Herbrandson et al. (2003a, b) report that if ingested, suspended solids may act as a secondary stressor by reducing food assimilation or increasing body weight and requiring cladocerans to expend more energy to stay at the desired depth within the water column. 4.4. B IOTIC

LIGAND MODEL

The BLM was developed to model metal speciation in diluents where metal binding may occur and predict concentrations needed to elicit mortality in certain organisms. The model also considers competition between the metal free-ion and cations such as Ca2+ , Mg2+ , Na+ , K+ , and H+ for binding sites on the biotic ligand (Di Toro

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et al., 2001; Santore et al., 2001). In tests with D. pulex, measured free-ion concentrations far exceeded the BLM predicted values, regardless of diluent. There was, however, better agreement between the BLM and measured free-ion concentrations for tests with D. clavipes, although measured values still exceeded the predicted values, particularly in tests conducted at 30 ◦ C. While studies investigating measured vs. BLM-predicted free-ion concentrations appear to be lacking, predicted metal speciation for the model is based on the Windermere humic aqueous model (WHAM) (Di Toro et al., 2001), which has been quite extensively evaluated. Tipping (1994), Vulkan et al. (2000), and DeSchamphelaere et al. (2005) found WHAM to accurately predict copper free-ion concentrations in waters with varying levels of organic carbon, although Vulkan et al. (2000) found WHAM only performed well if certain fulvic acid constants were adjusted. In contrast, Christensen et al. (1999) and Nolan et al. (2003) reported WHAM over-predicted organic carbon complexation of copper by an order of magnitude, resulting in measured free-ion concentrations exceeding predicted values. BLM predictions for total copper LC50 values were also made for D. pulex. In contrast to the free-ion results, the predicted LC50 values that were derived from the BLM usually fell within a factor of two of the actual values, which has been considered an adequate degree of comparability in other studies (Sciera et al., 2004; VanGerderen et al., 2005; Villavicencio et al., 2005). The model did overestimate toxicity for tests at 10 ◦ C in the wetland water and it underestimated toxicity in the soft laboratory water at 30 ◦ C, effects which could be due to the range of temperature calibration for the model, which is between 10 and 25 ◦ C (Di Toro et al., 2001; Santore et al., 2001). Sciera et al. (2004) and VanGerderen et al. (2005) reported that the biotic ligand model underestimated copper toxicity to fathead minnows in soft natural waters, with both suggesting that the model may not perform well for softer waters and that osmotic stress on the test organisms as they attempted to maintain sodium balance in soft waters may further influence the disparity. In contrast, Villavicencio et al. (2005) found the BLM to accurately predict LC50 values for three species of Daphnia in soft natural waters. Since the BLM has no parameter to account for binding with suspended solids, it was expected that predicted free-ion concentrations for the wetland water diluent would exceed the actual measured values. While this may partially explain the disparity for tests conducted at 10 ◦ C, the predictions for tests at 20 and 30 ◦ C were similar to values based on measured copper levels, and may indicate that TOC was more important as a complexing agent than TSS. Lu and Allen (2001) suggest that organic carbon can bind more copper than TSS, and they report an increase in TSS actually reduced the amount of copper bound to suspended solids due to interparticle interactions that decreased the number of metal binding sites. Erickson et al. (1996) also found that adding organic carbon to a diluent containing suspended solids could decrease the amount of copper bound by TSS.


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5. Conclusions Evaluating contaminant effects in habitats such as ephemeral wetlands provides an opportunity to examine the interaction between environment and a chemical stressor which is ultimately integrated through the response of the biota. Our data indicated that while this interaction does occur, it is not always straightforward. For example, while suspended and dissolved materials in the natural wetland water had a mostly ameliorative effect on copper toxicity to the invertebrate species tested, under certain temperature conditions, these materials may have enhanced sensitivity to the metal. There was also an indication that temperature optima exist for the test organisms, and that exposure to chemical stressors on either side of this range could influence their response. Comparison of acute effect levels based on total or free copper also support the concept that the free ion is a significant driver of toxicity to aquatic species. By examining effects based on free ion levels, it may be possible to reduce the influence of some factors that affect metal bioavailability (suspended solids) and focus on factors such as temperature which directly affect the physiology of the test organisms. An understanding of the effects that physicochemical parameters can have on metal toxicity to aquatic organisms can be facilitated through the use of models such as the BLM, although adjustments must be made for the factors that are most important in fluctuating environments such as ephemeral wetlands and for the species that occur in them.

Acknowledgements The authors would like to thank N. Cooper for her assistance in analyzing samples with atomic adsorption. The Ecotoxicology and Water Quality Research Laboratory at Oklahoma State University funded part of this research.

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