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a Dep. of Plant and Soil Sciences, Mississippi State Univ., Mississippi State, MS 39762
b Dep. of Wildlife and Fisheries, Mississippi State Univ., Mississippi State, MS 39762
c Dep. of Pharmacology, Univ. of Mississippi, University, MS 38677
Corresponding author (jhargreaves{at}CFR.MsState.Edu)
Received for publication March 13, 2000.
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ABSTRACT |
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 TOPNOTES ABSTRACT INTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Abbreviations: AmoFe, amorphous iron oxides • CARB, carbonate • CryFe, crystalline iron oxides • ERO, easily reducible oxides • EXC, exchangeable • OM, organic matter • RES, residual
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INTRODUCTION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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If the organism responsible for blue-green algae related off-flavor is identified in ponds containing fish near market size, one strategy to minimize the probability of occurrence of off-flavor is to apply an algicide to pond water. Currently, copper sulfate (CuSO4) is the sole USEPA-approved chemical for use in catfish ponds as an algicide. Channel catfish producers apply CuSO4 to ponds as needed throughout typical continuous production periods of 10 to 20 yr prior to pond draining and renovation.
Effective algicidal treatment requires CuSO4 · 5H2O application to ponds at 1% of the alkalinity (expressed as mg L-1 CaCO3) to obtain free Cu concentrations ranging from 0.25 to 3.0 mg L-1 (Gratzek, 1983). However, complete death of the phytoplankton bloom in a fish pond is undesirable as phytoplankton plays a critical role in maintaining suitable water quality for fish production. Regular, low-rate applications of CuSO4 may achieve a more desirable, partial reduction of a phytoplankton bloom (Tucker and van der Ploeg, 1999).
The long-term effects of periodic CuSO4 · 5H2O applications to catfish ponds are not known. As a first step, understanding the fractionation of Cu among solid-phase components can provide important fundamental information to assess the potential bioavailability and toxicity of Cu. Results of this study will provide a quantitative assessment of the effects of long-term, low-rate applications of CuSO4 · 5H2O. The purpose of this study was to investigate the accumulation, distribution, and potential bioavailability of Cu in the sediments of catfish ponds receiving weekly, low-rate applications of CuSO4 · 5H2O during the summer growing seasons over 3 yr.
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MATERIALS AND METHODS |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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Pond sediments were sampled following fish harvest and pond draining. The sampling rate necessary to obtain a representative sample from aquaculture pond sediments was based on recommendations by Ritvo et al. (1998). A composite (37 ha-1) sediment sample was collected using a uniform sampling grid with a 5-cm-diameter core tube from the sediment surface (0 to 5 cm) of each of eight nonamended and eight amended ponds for solid-phase Cu fractionation. In addition, surface (0 to 5 cm) sediment samples were collected using a uniform sampling grid at 74 ha-1 from one amended and one nonamended pond to determine spatial distribution of total Cu. Sediments were air-dried, ground with a ceramic mortar and pestle to pass a 2-mm mesh screen, and stored in sealed plastic bags until fractionation and toxicity testing.
Solid-Phase Copper Fractionation
Sediment Cu was separated into six operationally defined solid-phase fractions by selective, sequential dissolution. This method was based on the solubility of individual solid-phase components, and the selectivity and specificity of chemical reagents. The procedure provided a method to assess the gradient of physico–chemical association strength between Cu and various solid particles rather than actual speciation (Martin et al., 1987), thus providing a quantitative indication of the relative availability of Cu. The fractions are more likely to be operationally rather than chemically defined. However, each extractant in the selective, sequential procedure targeted one major solid-phase component. Therefore, the extractant-extractable metal is referred to by the targeted solid-phase component. In no case can an extractant remove all of a targeted solid-phase component without attacking some of the other components. In addition, redistribution and readsorption during sequential dissolution procedures may occur. Despite these shortcomings, common to any chemical extraction procedure, sequential dissolution techniques furnish more useful information on metal binding, mobility, and availability than can be obtained with a single extractant only.
The procedures employed in this study were based on those developed by Tessier et al. (1979) and Shuman (1985). Procedures for measurement of exchangeable, carbonate, residual, and total Cu were modified according to Han and Banin (1996)(1997). The procedure for measurement of the exchangeable fraction was modified by using a neutral nitrate salt (NH4NO3) as the extractant to avoid complexation with chloride and exchangeability of Mg that occurs with the original procedure. The procedure for measurement of the carbonate-bound fraction was modified to efficiently remove carbonate-bound metals from soils with varying carbonate content (Han and Banin, 1995). Finally, residual and total Cu were obtained by extraction with 4 M HNO3.
(i) Ammonium nitrate extractable—Soluble plus exchangeable copper (EXC). Twenty-five milliliters of 1 M NH4NO3 (pH adjusted to 7.0 with NH4OH) were added to 1.1 g of air-dry sediment (equivalent to 1 g of oven-dry sediment) in a 50-mL teflon centrifuge tube and the mixture was shaken for 30 min at 25°C and then centrifuged at 5000 x g for 10 min. The supernatant was decanted and filtered through a 0.45-µm filter. The sediment residue was retained for the next step. The same centrifugation–decantation procedure was used after each of the following extractions.
(ii) Sodium acetate/acetic aid extractable—Copper bound to carbonate (CARB). Twenty-five milliliters of 1 M NaOAc-HOAc at pH 5.0 were added to the residual sediment from step one. The mixture was shaken for 6 h.
(iii) Hydroxylamine hydrochloride extractable—Copper mostly bound to easily reducible oxides (ERO). Twenty-five milliliters of a 0.1 M NH2OH · HCl + 0.01 M HCl solution (pH 2) were added to the sediment residue from the previous step and shaken for 30 min. Metals extracted in this fraction are mostly from Mn oxides (Shuman, 1982). However, this acid can dissolve some Cu associated with organic matter, resulting in underestimation of organically bound Cu.
(iv) Hydrogen peroxide/ammonium nitrate extractable—Copper bound to organic matter (OM). Three milliliters of a 0.01 M HNO3 and 5 mL of 30% H2O2 were added to the sediment residue from the previous step and the mixture was digested in a water bath at 80°C for 2 h. An additional 2 mL of H2O2 were added and the mixture was heated for 1 h. Fifteen milliliters of 1 M NH4NO3 were then added and the sample was agitated for 10 min.
(v) Ammonium oxalate in the dark extractable—Copper bound to amorphous iron oxides (AmoFe). Twenty-five milliliters of a 0.2 M oxalate buffer solution (0.2 M (NH4)2C2O4–0.2 M H2C2O4 at pH 3.25) were added to the sediment residue from the previous step and the sample was shaken in the dark for 4 h (Shuman, 1982).
(vi) Hot hydroxylamine hydrochloride/acetic acid extractable—Copper bound to crystalline iron oxides (CryFe). Twenty-five milliliters of 0.04 M NH2OH · HCl in a 25% acetic acid solution were added to the sediment residue from the previous step and the sample was digested in a water bath at 97 to 100°C for 3 h.
(vii) Copper in the residual fraction (RES) and total copper. Twenty-five milliliters of 4 M HNO3 were added to the residue or sediment and the sample was transferred to a glass digestion tube. Digestion was conducted in a water bath at 80°C for 16 h (Sposito et al., 1982; Han and Banin, 1997). The tube was weighed before and after each step to calculate the remaining solution entrained in the residual sediment, which was subtracted from the measured metal concentration for each step.
In addition, water-soluble Cu was extracted using deionized water at a water to sediment ratio (w/w) of 2. Copper concentration in each extract was determined using atomic absorption spectroscopy. Free Cu ion in the water-soluble fraction was measured using a specific ion electrode.
Sediment Mineralogical Analysis
Samples were subjected to pretreatments to remove salts, carbonates, organic matter, and Fe oxides, and were then separated by sieving and centrifugation into sand-, silt-, and clay-sized fractions (Jackson, 1956; Dixon and White, 1997). Mineralogy of the clay-sized fraction of sediments was determined (Karathanasis and Hajek, 1982; Karathanasis and Harris, 1994). Clays were saturated with Mg and K salts, separately, for X-ray diffraction (XRD) and differential scanning calorimetric (DSC) analyses. A diffractometer (Philips X'Pert-MPD; Philips Electronics, Almelo, the Netherlands) equipped with a ceramic long, fine-focus Cu anode tube was used for XRD analysis. Samples were step-scanned from 2 to 30° 2
at 1 s per step with a step size of 0.03° 2. For DSC analysis (DSC 910S; TA Instruments, New Castle, DE), clay samples (Mg-saturated and dried at 54% relative humidity) were heated in covered aluminum pans from 5 to 625°C at 10°C per minute in an N2 atmosphere. An empty, covered aluminum pan was used as the reference (Karathanasis and Harris, 1994).
Data Analysis
The following parameters were used to describe the relative binding intensity of Cu in sediments (Banin et al., 1990; Han and Banin, 1997). The partition index (I) and the reduced partition index (IR) of Cu were calculated as follows:
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Comparisons of metal concentrations in the two sediments were made with t-tests. Significance levels of 0.05 and 0.01 were used in the analyses.
Sediment Toxicity Testing
Two model organisms, amphipods and common cattail, have been used previously to evaluate the toxicity of Cu in aquatic sediments (Huggett et al., 1999; Suedel et al., 1996; Deaver and Rodgers, 1996; Muller et al., 2001) and were used for toxicity testing in this study. Amphipods were cultured according to standard methods (USEPA, 1994) and 2- to 3-wk-old amphipods were used for bioassays, which consisted of 10-d static exposures (Nebeker et al., 1984). Cattail inflorescences were collected in the field, placed in plastic bags, transported to the laboratory, and stored at 20°C until testing. Prior to bioassays, viable seeds were separated from nonviable seeds gravimetrically (Moore et al., 1999) by placing part of the inflorescence in a blender with approximately 500 mL of tap water. Inflorescences were homogenized for 10 sec, whereupon floating debris was removed. Viable seeds used for testing were collected from the bottom of the blender. Cattail bioassays consisted of 7-d static exposures (Muller et al., 2001).
A 4:1 mixture (w/w) of water and each pond sediment sample was placed in glass borosilicate beakers. After settling of sediment particles, 10 amphipods and 10 cattail seeds were added to each respective test vessel. The toxicity of each sample was analyzed in three replicate test vessels. Amphipod tests were performed under constant aeration and amphipods were fed every other day with one drop of a solution made from yeast, alfalfa, and digested trout feed. At the end of the exposure period, amphipods were sieved, counted, and placed in 70% ethanol for growth determination. Cattail germination was determined by visual observation and the seedlings were gently placed in a 70% ethanol solution for root and shoot length measurement. Growth of amphipods and cattail roots and shoots after exposure was measured with a Videometric 150 image analyzer with software (American Innovision, San Diego, CA).
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RESULTS AND DISCUSSION |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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The spatial distribution of total Cu in nonamended and amended pond sediments was not uniform. Spatial variability of Cu concentrations in amended pond sediments was much greater than that in nonamended sediments. In nonamended ponds, sediment Cu concentrations ranged from 30 to 47 mg kg-1 (Fig. 1a). However, Cu concentrations in amended sediments ranged from <100 to >300 mg kg-1 (Fig. 1b), indicating that the fate of applied CuSO4 was not uniformly distributed in pond sediments and that composite sampling is essential for representative results. The greatest Cu concentration was measured in the sediment about 65 m down current from the point of CuSO4 · 5H2O application (Fig. 1b), suggesting that Cu precipitation from pond water was rapid.
Copper Fractionation in Sediments
Total Cu extracted by 4 M HNO3 was strongly correlated with the sum of the fractions by sequential dissolution. The proportion of total Cu extracted by 4 M HNO3 represented by the sum of the fractions obtained by sequential dissolution was 92.4% for nonamended sediment and 96.9% for Cu-amended sediment. These results indicate that the sequential dissolution technique resulted in excellent recovery of Cu from catfish pond sediments.
The fractionation of Cu was different between amended and nonamended pond sediments. In nonamended ponds, sediment Cu was mostly present in the AmoFe (28.3%), OM (21.4%), and CryFe (20.2%) fractions (Table 2). This indicates that Fe oxides or oxyhydroxides and organic matter play very important roles in binding Cu in nonamended sediments. Copper readily coprecipitates and forms solid solutions with Fe oxides and oxyhydroxides (Lindsay, 1979). The solubility of Cu oxide (tenorite) and hydroxide minerals is low and that of Cu carbonate is relatively high (Stumm and Morgan, 1981). However, Cu2+ may be chemisorbed or occluded in iron oxide coatings rather than as a separate phase through the formation of Cu2+–O–Fe3+ or Cu–O–Al bonds (McBride, 1981). McIntosh et al. (1978) reported that 50% of heavy metals were in the occluded phase of Fe and Mn oxides, and the concentration and distribution of Cu as well as Cd, Pb, Ni, and Zn were positively correlated to the amount of hydrous iron and manganese oxides in the sediments. These observations support the observation that Cu in nonamended sediments was distributed primarily in fractions associated with Fe (i.e., AmoFe and CryFe fractions).
In contrast to nonamended pond sediments, sediment Cu in CuSO4–amended ponds was mainly present in the CARB (31.6%) and OM (31.1%) fractions. There was also a large proportion of Cu in the AmoFe (22.1%) fraction. Water used to fill the ponds sampled in this study had moderately high alkalinity (100 to 200 mg L-1 as CaCO3) that would promote precipitation of CuCO3 despite the relatively high solubility of this compound. The association of excess soluble Cu added to soils with the carbonate phase tends to be temporary. CuCO3 is not very stable in aqueous alkaline environments due to larger ionic size differences between Cu (0.72 x 10-10 m) and Ca (0.99 x 10-10 m) (McBride, 1981) and the relatively high solubility of CuCO3 (Ksp = 1.4 x 10-10) (Dean, 1973). Thus, following initial precipitation, Cu in CuSO4–amended pond sediments may redistribute with time from the CARB fraction into the AmoFe-bound and OM fractions. Han and Banin (1997)(1999) reported that added soluble Cu was redistributed from the EXC and CARB fraction into the ERO and OM fractions in arid-zone soils. However, Cu2+ may substitute freely for Mg in magnesium carbonate or Fe in iron carbonate (McBride, 1981).
The large proportion of Cu in the OM fraction can be related to the uptake by and subsequent sedimentation of dead phytoplankton. Additionally, Cu can be adsorbed to organic matter, particularly humic substances. Copper can form the core of a very stable inner sphere complex with humic substances (McBride, 1981).
Copper extracted from many river and lake sediments is primarily in the residual fraction, followed by the organic matter–bound fraction. In two Quebec river sediments, Cu was mostly in the residual (38%), organic matter–bound (25%), Fe and Mn oxide–bound (20%), and carbonate-bound (15%) fractions (Tessier et al., 1979). Copper was mainly in the residual (53%) and organic matter–bound (30%) fractions in sediments of the Gulf Intracoastal Waterway near Galveston, Texas (Lindau and Hossner, 1982). These results indicate that Cu will ultimately redistribute to fractions of low bioavailability. Copper in CuSO4–amended catfish pond sediments was present in more labile fractions than in many river and lake sediments, reflecting a history of recent CuSO4 additions. Additional research is required to assess the kinetics of Cu redistribution in aquatic sediments.
The relative binding intensity (IR) of Cu in nonamended sediments was 1.5 times greater than that of Cu in CuSO4–amended sediments, indicating that accumulation of Cu in amended sediments over the time frame (3 yr) of CuSO4 · 5H2O applications occurred in fractions with potentially greater bioavailability. The greatest proportional increases of Cu in amended pond sediments occurred in the EXC and CARB fractions (Table 2). As discussed previously, IR can range from 0.02 to 1. The value of IR was strongly related to the total inputs of Cu, sediment properties, and time scale (Table 2). Han and Banin (1997)(1999) reported that the IR of Cu was inversely related to Cu loading in arid-zone soils, and IR in a loessal soil was higher than that in a sandy soil. Binding intensity (as indicated by IR) can be expected to increase with time as a result of redistribution of Cu from more labile into more stable fractions. McLaren and Ritchie (1993) reported that a high proportion of the applied Cu as copper sulfate in a lateritic sandy soil in western Australia was initially associated with the soil organic matter. During the course of the trial (20 yr), a substantial proportion of Cu redistributed to the residual fraction.
Solution Copper and Availability of Copper in Sediments
Copper concentrations in the soluble and EXC fractions in amended sediments were much higher than those in nonamended sediments (Table 2). Water-soluble and EXC Cu in the amended sediments were 3 and 11 times higher than those in nonamended sediments, respectively. This indicates that the potential bioavailability of Cu in amended pond sediments was greater than that in nonamended sediments.
In the water-soluble fraction, no free Cu ion was detected by specific ion electrode (detection limit = 10-8 M; Orion Research, 1996) in extracts from both nonamended and amended sediments (data not shown), implying that Cu in the water-soluble fraction was in complexes formed with inorganic and organic ligands. Thus, the availability of Cu in the amended pond sediment systems may be relatively low.
Exchangeable Cu is the key fraction bridging solid-phase and solution components. There was a strong correlation between water-soluble and EXC Cu in sediment from both treatments (Table 3). Bioavailability, as indicated by water-soluble and EXC Cu concentrations, was strongly controlled by solid-phase components in nonamended sediments. Copper concentrations in water-soluble and EXC extracts of nonamended pond sediments were strongly correlated with sediment solid-phase Cu concentrations in CARB, AmoFe, CryFe, or OM fractions (Table 3). In contrast, Cu concentrations in water-soluble and EXC extracts of amended pond sediments were not correlated with any solid-phase Cu fraction concentrations (Table 3). These results suggest that the bioavailability of Cu is directly related to solid-phase Cu concentrations at low solid-phase Cu concentrations, but increases in bioavailability are not directly related to solid-phase Cu concentrations above certain threshold solid-phase Cu concentrations at relatively high solid-phase Cu concentrations.
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63 µg Cu L-1 may cause effects to amphipods. Nonetheless, IWTU may underestimate Cu toxicity because interstitial water is not the only route of exposure. Amphipods feed on detrital particles at the sediment–water interface (Chapman et al., 1998). Dietary exposure to Cu may be enhanced by the consumption of settled phytoplankton killed through Cu uptake. In addition, organisms may be exposed to Cu by the diffusion of free Cu from sediments into the overlying water. Suedel et al. (1996) observed a relationship between the overlying water Cu concentrations and toxicity to amphipods in Cu-amended sediment bioassays.
In this study, sediments were air-dried following collection for the Cu fractionation study and then rewetted for toxicity testing. The bioavailability of Cu may have been altered (reduced) by precipitation of Cu oxides and carbonates upon drying. Further studies are required to evaluate the toxicity of freshly collected sediment, which should include measurement of porewater Cu concentrations. Additional research that relates sediment Cu concentration determined from air-dried samples to interstitial free Cu ion concentration is also required. Finally, the effect of sediment Cu concentration on microbial activity in catfish ponds is not known. Inhibition of microbial activity could have profound effects on catfish pond oxygen dynamics and the regeneration of nutrients from organic matter mineralization.
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NOTES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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TOPNOTESABSTRACTINTRODUCTIONMATERIALS AND METHODSRESULTS AND DISCUSSIONREFERENCES |
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