COPPER

RETENTION

AND

TOXICITY

IN A FRESHWATER

SEDIMENT

C. A. F L E M M I N G and J. T. TREVORS* Department of Environmental Biology, University of Guelph, Guelph, Ontario, Canada NI G 2 W1

(Received February 29, 1988; revised June 9, 1988) Abstract. The effects of Cu(II) sulfate on sediment respiration were investigated in a 3-phase aquatic microcosm, containing a calcareous, southern Ontario stream sediment. In Cu2+ treated flask-microcosms, with the pH restored to 7.1, both aerobic and anaerobic CO2 evolution were unaffected by 5000 lig Cu g- 1 sediment over a 40-day period at 15 °C. Oxygen consumption in sediment was initially unaffected by 5000 ~tgCu g- 1. However, after 35 to 40 days, a significant reduction of 28% was observed. The added Cu2. was removed from the water column and the sediment solution. In microcosms containing 5000 ~tgg- 1 of total Cu, only 1.00 + 0.76 ~tgg- 1 was water soluble Cu, and the free cupric cation (Cu2+ ) concentration was below the detection limit of the specific ion electrode (less than 0.01 ~tgg- ~). Maximum Cu retention (98.6%) was observed at 2800 ~tg Cu g- 1, above which fractional retention decreased. In a calcareous, organic rich, sediment of pH 7.1, Cu2+ was essentially unvailable to exert a toxic effect on respiration.

1. Introduction C o p p e r levels can be elevated in soils and sediments in localized geographic areas due to the smelting of ores (Beavington, 1973, 1977), mining operations (Lopez and Lee, 1977), industrial and domestic waste emission ( F o r s t n e r a n d W i t t m a n n , 1979), sewage sludge applications to soil (Barkay et al., 1985), and by its w i d e s p r e a d application to water systems as a molluscicide (Cheng, 1979) and algicide ( M a c K e n t h u n a n d Cooley, 1952; Elder and H o m e , 1978). C o p p e r entering freshwater systems is rapidly r e m o v e d to the sediment as a result o f retention by particulate matter and precipitation (Jackson, 1978). It is i m p o r t a n t to assess a n d predict the effects that accumulated Cu m a y exert on indigenous sediment microbial communities because microorganisms play an essential role in the cycling o f elements (C, N, P, S), regeneration o f inorganic nutrients, and nutrient transformations in the environment. Some studies have examined the toxicity o f metals, including Cu, on microbial activities in the natural environment (Beavington 1973; Babich and Stotzky, 1985). M o s t studies report only the total metal levels in a soil or sediment and provide little information about the physico-chemical nature o f the sediment or soil samples. The present study was u n d e r t a k e n to assess Cu a ÷ chemistry and toxicity in a well-characterized sediment.

* Author for all correspondence. Water, Air, and Soil Pollution 40 (1988) 419-432. © 1988 bv Kluwer Academic Publishers.

420

C. A. FLEMMING AND J. T. TREVORS

2. Materials and Methods 2.1.

PHYSICAL AND CHEMICAL ANALYSIS

Sediment and water samples were collected from the east branch of the Canagagigue Creek, Floradale, Ontario, Canada, an area described by Dance et aL (1979). Sediment pH and Eh were measured using standard methods (Rodina, 1972). Sediment particle fractionation was determined using the hydrometer method (Day, 1965). Organic matter and organic C were determined by loss on ignition (Ball, 1964). Total C and inorganic C were estimated using a Leco induction furnace equipped with a Leco carbon determinator (Laboratory Equipment Corporation, Model WR12). Total organic N was determined by the Kjeldahl method of Bremner and Muhaney (1982). Sediment mineralogy was determined semiquantitatively by X-ray diffraction (XRD) analysis of random powder sediment samples on a diffractometer (Ragaku Geigerflex) employing CuK radiation. Samples were rotated from 3 to 75 ° 2-theta. Cation exchange capacity (CEC) of sediment was determined outlined by Thomas (1982), and expressed as meq 100 g - 1 dry sediment. Calcium, Mg, K, and Na were determined using an atomic absorption spectrophotometer (Varian Techtron AA6) as recommended by the manufacturer. Total P and N were determined colorimetrically using molybdate blue and endophenol color development, respectively, on a Technicon autoanalyzer (Technicon, 1978). Dissolved organic carbon (DOC) was determined in both sediment solution and stream water samples. Water soluble C was extracted from wet sediment (1 : 5 dilution with distilled water) by shaking at 300 rpm for I hr followed by centrifugation at 3000 x g for 20 min. Filtered sediment (Whatman No. 42) extract and filtered stream water samples were analyzed colorimetrically using an autoanalyser for organic C (Method No. 455-76W/A; Technicon, 1978). Aqueous potassium biphthalate was used as the organic C standard. Nitrate (NO;-) in the sediment extract and stream water was determined colorimetrically (at 420 nm) using the brucine-sulfanilic acid method (Trevors, 1984). Carbonate alkalinity was estimated from measured values of the partial pressure of CO2 (pCO2) and pH, as described by Sunda and Hanson (1979). Carbon dioxide was determined in the gas phase as described in Section 2.2. Calculations used the dissociation constants reported by Manahan (1984) and Weast (1980). Electrical conductivity (EC) of sediment solution extracts and stream water was calculated from resistance measurements taken with an ohm meter. Specific conductance was calculated using the method of Rainwater and Thatcher (1960). The ionic strength (I) of samples was calculated from EC values using the equation I = EC x 0.013 (Lindsay, 1979). Sulfate levels (SO 4 2)were estimated from EC values (Wetzel, 1975). 2.2.

SEDIMENT RESPIRATION

Oxygen and CO 2 concentrations were measured in the gas phase of 3-phase flaskmicrocosms (sediment/water column/gas phase) using thermal conductivity gas chro-

COPPER RETENTION AND TOXICITY IN A FRESHWATER SEDIMENT

421

matography as described by Trevors (1983). Flasks (50 mL) contained 10.0 g wet sediment and 5 mL of stream water. Filter-sterilized CuSO4- 5HzO stock solutions were added to give Cu levels ranging from 0 to 5000/~g g - i dry weight sediment. Equivalent amounts of sterile NazSO 4 stock solution were added to another series of flask-microcosms which served as sulfate treated controls. After addition, flasks were shaken at 100 rpm for 15 rain, and sealed with serum stoppers. Flasks were made anaerobic by repeated evacuation (70 cm Hg vacuum) and refilling with pure He until the 0 2 concentration in the gas phase was less than 0.5~o. Sediment pH and Eh were measured after CuSO 4 addition, as CuSO4 caused a decrease in pH which was naturally neutralized 7 to 9 days later. In some experiments, flasks were sealed only after the pH was restored to 7.1. Sterile sediment samples were prepared by autoclaving sediment for 1 hr on 3 consecutive days, with 30 °C incubation for 1 day between autoclaving. Filtersterilized CuSO4 and stream water were added after autoclaving. All experiments were done in triplicate with static incubation at 15 °C in the dark. This temperature was chosen to approximate the temperature measured at sampling times throughout the summer months (Table I).

TABLE I Selected sediment characteristics Texture ( ~ ) sand silt clay

70.0 + 4.0 21.0 +_ 2.7 9.0 +0.1

Dry weight (~o wet weight)

55.0 _+ 4.5

Total C (~o) Loss on ignition ( ~ ) Inorganic C (~o) Total N (~o) C : N ratio CEC (meg. 100 g - ~) p H (in H 2 0 ) (in CaClz) Eh (mV)

4.5 6.9 1.7 0.31 16:1 48.7 7.3 6.9 - 67

+_ 0.21 + 0.25 + 0.05 + 0.002 + 1.2 + 0.1 + 0.1

Temperature ( ° C at time of sampling)

15.8 + 2.0

Mineralogy

quartz dolomite calcite 1.1 + 106 ___0.14 x 106 g - 1

Aerobic heterotrophs (CFU)

Values are given as the m e a n + S.D. (n = 2). CEC cation exchange capacity (Ca, Mg, K, and Na). The streamwater had a pH at the sediment interface of 7.1 + 0.21 and a water column p H of 8.1 + 0.28. Total hardness (mg L - x CaCO2) was 225" 0 + 21.2.

422 2.3.

C. A. FLEMM1NG AND J. T. TREVORS E N U M E R A T I O N OF M I C R O O R G A N I S M S

Aerobic, heterotrophic bacterial numbers were estimated by plating serial dilutions of sediment samples on Nutrient Agar (Difco). Triplicate plates from each dilution were incubated at 15 °C in the dark for up to 5 weeks. The extended incubation time was necessary as some colonies were slow growing, and it was observed that colony-forming units were increasing up to five weeks' incubation. 2.4. Cu

R E T E N T I O N BY S E D I M E N T

One g of air-dried, sieved (less than 2 ram) sediment was suspended in 40 mL of the filter-sterilized CuSO4 solution in a 50 mL Erlenmeyer flask. Cu 2+ concentrations ranged from 0 to 250 ~tg Cu 2÷ m L - 1 (0 to 10,000 I-tgg - 1 dry weight sediment). Flasks were shaken at 250 rpm for 4 hr at 15 ° C and allowed to settle for 2 hr. Total equilibrium time was 6 hr. Suspensions were filtered through Whatman No. 42 paper. Equilibrium concentrations of total dissolved Cu and Cu 2 ÷ in solution were determined using atomic absorption spectrophotometry and a Cu 2÷ specific ion electrode (Orion Research), Cambridge, MA, U.S.A.), respectively. 2.5. STATISTICAL ANALYSES Data were analysed using a statistical package (Ed. Sci., Modesto, California) on an Apple II plus microcomputer. Analysis of variance (ANOVA) was done at (P = 0.05), followed by a Student Newman-Keuls (SNK) multiple range test (P = 0.05) if significant Fvalues were obtained. Respiration data were analysed by comparing data points for treatments and controls at three times during the incubation period: 2 days, 12 to 17 days, and 29 to 34 days. 3. Results and Discussion Chemical and physical characterization of sediment and stream water samples was important as these parameters control speciation and toxicity of Cu in the environment, as well as influence microbial communities. Selected sediment and stream water characteristics are summarized in Tables I and II. The mean p H of the sediment and stream water was 7.1 and 8.1, respectively (Table I). The sampling site is a hard-water stream, with an average mean total hardness of 225 mg L - 1 CaCO3. High water hardness significantly influences Cu speciation and causes precipitation, thereby, lowering Cu availability and toxicity in the environment (Stiff, 1971; Hodson et aL, 1979). The average number of aerobic sediment heterotrophic microorganisms was 1.1 × 10 6 g - 1 dry weight. Anaerobic microorganisms were not enumerated, but can be as numerous as aerobic microorganisms in sediment (Parkes et al., 1979). The sediment had a sandy loam texture with an average organic matter content of 6.9%. Semi-quantitative X-ray diffraction data showed dolomite (CaCO3" MgCO3) and calcite (CaCO3) in near aqual proportions in the sediment (Table I). The presence of these minerals can significantly decrease the free Cu 2 + concentration and also buffer the streamwater.

COPPERRETENTIONANDTOXICITYIN A FRESHWATERSEDIMENT

423

TABLE II Concentrations ( - logM) of selected elements and ligand in sediment solution and stream water, and other solution chemical data Concentration ( - logM) Sediment solution

Stream water

Elemems Ca Mg Na K

2.9 3.4 1.4 4.4

2.6 2.9 3.1 4.3

3.4 4.7 3.2 3.2 nil 873.0

2.1 5.2 3.5 3.1 2.3 64.4

Ligands CO3 PO2 NO3 SO4 CI Dissolved organic C (rag L- 1) Electrical conductivity(EC) (mmho cm- 1)

0.91

0.75

Ionic strength (I) (M) - log(PCO2)a

0.012 2.8

0.010 1.7

a PCO2

=

partial pressure of C O 2 in atmospheres.

Total sediment N was O . 3 1 ~ with a C : N ratio of 16 : 1 (Table I). The concentrations of selected metals and ligands which can influence Cu speciation are summarized in Table II, along with other selected chemical characteristics of the sediment solution and stream water. These chemical parameters are important when studying environmental samples because they dictate metal toxicity and retention. For example, the sediment Eh increased to + 317 mV from - 67 mV, when exposed to 5000 gg Cu g - 1. Oxygen consumption and CO2 evolution were measured in the headspace of flaskmicrocosms over a 30 to 40 day period to assay long-term sediment respiration in the presence of added CuSO4. Sodium sulfate (Na2SO4) controls confirmed that the S O l anion had no effect on 0 2 consumption or CO 2 evolution at S O l - levels identical to those used in CuSO4 treatment experiments. Respiration was examined in flask-microcosms sealed immediately, and sealed 7 to 9 days after CuSO4 addition. This was done because CuSO 4 addition caused a pH decrease which was neutralized after 7 to 9 days (Figure 1) due to the buffering capacity of the sediment. This allowed the combined effects of pH and Cu to be differentiated from the effect of Cu alone. Flask-microcosms sealed immediately after the addition of CuSO4 gave different results for respiration than flask-microcosms allowed to return to their original pH of 7.1. For this reason, both sets of data are presented (Figure 2A and 2B). There were no significant differences in initial 02 consumption analyzed at both day 2 and day 14. However, overall 02 consumption after 40 days was higher in

424

C. A. F L E M M I N G AND J. T. TREVORS

7.5

7.05

o

6. pH

• •

0

2

4

6 Days

8

10

Fig. 1. Sediment pH plotted against time after C u S O 4 addition. Sediment pH was measured in flasks after addition of 1000 (A), 2500 ( A ) and 5000 gg Cu g - 1 (o). Initial sediment pH was 7.1 prior to addition of CHSO 4 (o).

control flask-microcosms (sealed 9 days after Cu addition) than in those treated with 500 to 5000 gg Cu g - 1 (Figure 2B). Aerobic CO 2 evolution in flasks sealed immediately increased significantly (day 2) at treatment levels of 2500 and 5000 gg Cu g - 1. At day 14, significant differences in CO2 evolution between controls and high Cu 2 ÷ treated flasks-microcosms were observed. However, after 36 days, there were no significant differences (Figure 3A). Flask-microcosms sealed 9 days after Cu addition (Figure 3B) displayed no significant differences in initial CO 2 evolution (day 2), at 2 weeks and after 36 days. Control flask-microcosms had slightly (not statistically significant) higher CO 2 production than Cu-treated samples. In flasks sealed at variable pHs, the initial anaerobic CO2 evolution (at day 2 and 7) in flask-microcosms treated with 1000 and 5000 Ixg Cu 2+ g - 1, were significantly higher than in the controls (Figure 4A). Total CO2 evolution was not different in treatment and control flask-microcosms at day 39 (Figure 4B). Anaerobic flask-microcosms sealed at a neutral pH (Figure 4B) showed no significant differences in CO 2 evolution troughout the experiment. These results were similar to aerobically produced CO 2 in unsealed flask-microcosms. Total CO 2 evolution in anaerobically incubated sediment was about 6 gmol CO2g -1 dry weight sediment (Figure4); 2-fold less than in aerobically incubated sediment. This is normal as anaerobic respiration is less efficient than aerobic respiration.

80

425

A

60 ICn

E

x

40

×

i[

E m

g 20

t~

I'M

C)

I

0

8

16

I

I

24

I

, I

32

40

Days 8O

+0

B

T

+

y

E 40 "0

B m

g 2O u

0

'

8"

'

' 16

'

21 4

'

' 32

'

' 40

Days Fig. 2. Effect of C u S O 4 o n sediment O 2 consumption in aerobically incubated flask-microcosms. Flasks were sealed immediately (A) and 9 days after CuSO4 addition (i.e,, after pH had returned to original value ofT.1 in all flasks (B). CuSO 4 was added to sediment at 0 (o), 100 (o), 500 ( A ) , 1000 ( A ) , 2500 (V]), and 5000 gg Cu g - t ( I ) , Mean values + S.D. (n = 3) are plotted. The dashed line ( - . - ) indicates 100~o 0 2 consumed in the gas phase. Points labelled with different letters are significantly different (P = 0.05) as determined with S N K test.

426

C. A. F L E M M I N G A N D J. T. T R E V O R S

't~ 40

>

2O

r,4

0

8

16 Days

24

32

/,0

B

20 c.,4 I

I

8

I

I

I

16

I

24

!

I

32

I

I

40

Days Fig. 3. Effect of C H S O 4 o n CO 2 evolution in aerobically incubated flask-microcosms. Flasks were sealed immediately (A) and 9 days after CuSO4 addition (B). Symbols same as in Figure 2.

Initial CO/evolution (day 2) in flask-microcosms sealed immediately increased with increased Cu concentrations (Figures 3 and 4). This significant increase at high Cu levels is thought to be a chemical effect due to the decrease in pH, because it was absent in flask-microcosms sealed after pH neutralization (Figure 3 and 4). The effect of a pH decrease on CO 2 evolution was examined in sterile sediment (Figure 5). As the sediment pH decreased below 6.0, the initial CO 2 evolution (day 2) increased. This CO 2 production was due to the formation of carbonic acid (H2CO3) from carbonate minerals in sediment dissolving at acidic pHs. Carbon dioxide was liberated as the sediment pH was neutralized. A drop of H 2 S O 4 added to the sediment sample caused violent effervescing (CO 2 evolution), indicative of the presence of carbonate minerals. Sediment respiration was also examined in autoclaved, control sediment samples. Total 02 consumption after 40 days was 14.6~ of non-sterile controls. Total CO 2

427

COPPER RETENTION AND TOXICITY IN A FRESHWATER S E D I M E N T

16

A

A

12

c]J

c_~ 4

i

i

i

16

i

24

i

i

s ,

2

J

40

Days 16

B

--12 O E -.) v

(U

>8 O

x

O C_D

4

Days Fig. 4. Effect of C u S O 4 o n CO 2 evolution in anaerobically incubated flask-microcosms. (A) Flasks were sealed and immediately made anaerobic, followed by treatment with CuSO 4. (B) Flasks were treated with CuSO4, incubated 9 days and then sealed and made anaerobic. Symbols same as in Figure 2.

428

C. A. F L E M M I N G A N D J. T. T R E V O R S

120

100

80

"7 -O

60

E

o--.~

>

'~ 40 Or',J

20

0

4"0

~ 5.0

6-0

7.0

pH Fig. 5.

Effect o f p H oi1 CO 2 evolution in autoclaved flask-microcosms. Sediment pH was initially 7.1. CO 2 levels were measured in the gas phase 2 days after pH perturbation.

evolution was 5.5 % of that reported in non-sterile sediment (data not shown). This indicated that the respiration rates observed in this study can be attributed mainly to biological respiration, except where carbonic acid was formed. Two months after CuSO 4 was added to flask-microcosms, viable aerobic heterotrophic bacteria from the sediment were enumerated on Nutrient Agar (Table IV). There were no significant differences between colony forming units in control and sediment samples exposed to concentrations up to 1000 ~tg Cu g - 1. However, a 2-fold increase in colony forming units was observed in sediment samples treated with 2550 and 5000 ~tg Cu g - 1. It is thought that increased numbers were due to the development of a population of Cu-tolerant microorganisms. When the control sediment dilutions were

COPPER RETENTION AND TOXICITY IN A FRESHWATER SEDIMENT

429

100 80

=

o .El

60

0

.-z 40

20

0

2000

& 000

6 000

8000

Total Cu(pg g-t) Fig. 6. Copper adsorption by sediment at increasing Cu concentrations. Sediment-bound Cu was determined as the difference between equilibrium and total levels of Cu in solution after 6 h, as measured using an atomic absorption spectrophotometer. Mean values for n = 2 are plotted.

TABLE III Effect of a 2-too exposure to different Cu concentrations on numbers of colony-forming units (CFU) of heterotrophic bacteria in sediment Cu (~tg g -

1)

0 (control) 100 500 1000 2500 5000

CFU g - l + S.D. x l0 s (n = 3) 11.17 9.90 11.81 9.72 22.34 20.49

+ 2.61 _+ 2.56 _+ 1.38 + 3.13 + 2.76 a _+ 2.41 a

Statistically significant difference at P = 0.05 as determined by the SNK test.

I0000

430

¢. A. FLEMMING AND J. T. TREVORS

plated on Nutrient Agar amended with 500 gg Cu g - 1, only 0.15~o of the total colony forming units were able to form colonies. However, in sediment previously exposed to 5000 lag Cu g - 1, 42.4~ of the colony forming units were viable on the agar containing Cu. This represented a 500-fold increase in Cu-tolerance over that observed in the control sediment (data not shown). Additional experiments revealed that equivalent concentrations of Na 2 SO 4 in agar had no effect on the number of colony forming units (data not shown). Interpretation of the respiration data is highly dependent on the chemical physical and biological state of samples. The sediment pH initially decreased from pH 7.1 to as low as 5.0. It is doubtful that this pH drop influenced sediment microbial activities directly. Effects of acidification have been extensively reported in the literature. Acidification of sediment to pH 4.2 in microcosms had no effect on anaerobic CO 2 evolution and CH 4 production (Furutani et al., 1984). In soil samples. CO2 evolution was not affected at pH 3.0 (Bewley and Stotzky, 1983). Litter decomposition appears to be reduced in acidic lakes (Dillon et aL, 1983), but a pH less than 5.0 had no influence on litter colonization, adenosine triphosphate (ATP) biomass or cellulose mineralization (McKinley and Vestal, 1982). Cu retention by sediment was examined after 6 hr for initial Cu levels ranging from 0 to 10 000 lag g - 1 (Figure 6). Retained Cu was calculated by subtracting dissolved Cu at equilibrium from the initial concentrations of Cu in solution. It is noteworthy that Cu retention refers to Cu removed from solution; this involves the process of sorption, chelation, reduction, and precipitation, which are indistinguishable here. Precipitate formation will be significant at pH 7.0 (Pickering, 1979; Ferrah and Pickering, 1978; Sanches and Lee, 1973; Kishk etaL, 1973). Copper retention increased in a linear manner with increased Cu added, up to 2800 lag g - 1 with a slope of 0.97. The curve deviated slightly at 5000 gg Cu g - 1 and gradually levelled off at 10 000 lag Cu g - a, as the maximum sediment retention capacity was approached. The maximum percent Cu retained 98.6~ occurred at 2800 lag Cu 2 ÷ g - 1 addition; above which the fractional retention of Cu in sediments decreased. Water soluble Cu in the sediment solution of Cu-treated microcosms was low (less than 1.2 lag mL - 1). Free Cu 2 ÷ (less than 0.01 lag g - 1) was not detectable with a Cu 2 + specific electrode. Computer modeling of Cu e ÷ speciation in the stream water and sediment solution was done using the program G E O C H E M (Mattigod and Sposito, 1979), and the chemical data in Table II. G E O C H E M predicted that 87.5 ~o of dissolved Cu was in the form of carbonate complexes. 12~o complexed with dissolved organic matter, 0.3 ~ in the form of hydroxide complexes and only 0.2~ was in the free ionic form. The concentrations of Cu reported to be toxic in natural environments such as soils, sediments and waters, varies widely in the literature. Toxicity is largely dependent upon availability of Cu e+ (Zevenhuizen etaL, 1979) which is dependent upon specific physico-chemical characteristics that influence the speciation of the metal. Many of these characteristics were measured in the present study, thus allowing a better understanding of Cu toxicity. In many studies, metal toxicity is reported as the toxicity related

COPPER RETENTIONAND TOXICITYIN A FRESHWATERSEDIMENT

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to total metal added, rather than metal available or ionic metal concentrations. In the present study, CuSO4 had little effect on microbial respiration even at elevated levels. High Cu 2 + concentrations were rapidly retained by components of the sediment, and complexed and/or precipitated, thus decreasing Cu toxicity to sediment respiration.

Acknowledgments C.A.F. was supported by a NSERC (Canada) postgraduate scholarship. J.T.T. is grateful to the NSERC operating grants program for financial support of the project. Sincere appreciation is expressed to B. McGavin for typing the manuscript, and Dr R. Simard for assistance with the GEOCHEM program.

References Babich, H. and Stotzky, G.: 1985, Environ. Res. 36, 111. Babich, H. and Stotzky, G.: 1980, CRC Crit. Rev. Microbiol. 8, 99. Ball, D. F.: 1964, J. Soil Sci. 15, 84. Barkay, T., Tripp, S. C., and Olson, B. H.: 1985, Appl. Environ. MicrobioL 49, 333. Beavington, F.: 1973, Aust. J. Soil. Res.ll, 27. Beavington, F.: 1977, Environ. Pollut. 13, 127. Bewley, R. J. F. and Stotzky, G.: 1983, Soil Biol. Biochem. 15, 425. Bremner, J. M. and Mulvaney, C. S.: 1982, in A. L. Page, R. H. Miller, D. R. Keeney (eds.),Methods of Soil Analysis, Am. Soc. Agron. Inc., Madison, WI., p. 395. Cheng, T. C.: 1979, in J.O. Nriagu (ed.), Copper in the Environment, Jhn Wiley and Sons, Toronto pp. 401-432. Dance, K. W., Hynes, H. B. N., and Kaushik, N. B.: 1979, Arch. Hydrobiot. 87, 139. Day, P. R.: 1965, in C. A. Black (ed.), Methods of SoilAnalysis, Am. Soc. Agron. Inc., Madison, WI., p. 545. Dillon, P. J., Yan, N. D., and Harvey, H. H.: 1983, CRC Crit. Rev. Environ. Cont. 13, 167. Elder, J. F. and Home, A. J.: 1978, Environ Management. 2, 17. Ferrah, H. and Picketing, W. F.: 1978, Water, Air, and Soil. Pollut. 9, 491. Forstner, U. and Wittman, G. T. W.: 1979, Metal Pollution in the Aquatic Environment, Springer-Verlag, Berlin, pp. 1-486. Furutani, A., Rudd, J. W. M., and Kelly, C. A.: 1984, Can. or. MicrobioL 30, 1408. Hodson, P. V., Borgmann, U., and Shear, H.: 1979, in J. O. Nriagu (ed.), Copper in the Environment. John Wiely and Sons, Toronto, p. 307. Jackson, T. A.: 1978, Environ. Geol. 2, 173. Kishk, F. M., Hassan, M. N., Ghanem, I., and El-Sissy, L.: 1973, 1"1. Soil 39, 487. Lindsay, W. L.: 1979, Chemical Equilibria in Soils, John Wiley and Sons, Inc., New York, p. 449. Lopez, J. M.. and Lee, G. F.: 1977, Water, Air and SoitPoillut. 8, 373. MacKenthun, K. M. and Cooley, H. L.: 1952, Trans. Wis. Acad. Sci. Arts Letts. 41, 177. Manahan, S. E.: 1984, Environmental Chemistry, 4th ed., Willard Grant Press, Boston, p. 612. Mattigod, S. V. and Sposito, G.: 1979, in E. A. Jenne (ed.), ChemicalModeling in Aqueous Systems: Speciation, Sorption, Solubility and Kinetics, ASC Symposium Series, Vol. 93, p. 837, American Chemical Society, Washington, D.C. McKinley, V. L. and Vestal, J. R.: 1982, Appl. Environ. MicrobioL 43, 1188. Parkes, R. J., Bryder, M. J., Maden, R. H., and Poole, H. J.: 1979, in C. D. Litchfield and P. L. Seyfried (eds.), Methodology for Biomass Determinations and Microbial Activities in Sediments, Amer. Soc. Test. Mat. Philadelphia, p. 193. Picketing, W. F.: 1979, in J. Nriagu (ed.), Copper in the Environment, John Wiley and Sons Inc., New York, p. 217. Rainwater, F. H. and Thatcher, L. L.: 1960, Methodsfor Collection and Analysis of Water Samples, U.S. Geol. Surv. Water Supply Paper No. 1454, p. 301.

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Rodina, A. G.: 1972, Methods in Aquatic Microbiology, University Park Press, Baltimore, p. 461; Sanchez, I. and Lee G. F.: 1978, Water Res. 12, 899. Stiff, M. J.: 1971, Water Res. 5, 171. Sundra, W. G. and Hanson, P. J.: 1979, ACS Symposium Series 93, 149. Technicon: 1978, Technicon AAII S.C. Colourimetrie Technicon Industrial Systems, Industrial Methods No. 455-76 W/A. Tarrytown, NY. Thomas, G. W.: 1982, in A. L. Page, R. H. Miller, and D. R. Keeney (eds.), Methods of SoilAnalyses, Am. Soc. Agron. Inc., Madison, WI, p. 159. Trevors, J. T.: 1984, Soil Microbiology Laboratory Exercises and Supplemental Lecture Notes, University of Guelph, Guelph, Canada, p. 96. Trevors, J. T.: 1983, Biotechnol. Letts. 5, 625. Varian Teehtron: 1981, Varian Instruments at Work. No. AA-12, Operation Manual for Flame/CRA-90. Methods for Flame Analysis (AA-6). Varian Techtron. Weast, R. C.: 1980, C.R.C. Handbook of Chemistry and Physics, CRC Press Inc., F1. Wetzel R. G.: 1975, Limnology, Saunders College Publishing, Philadelphia, PA, p. 743. Zevenhuizen, L. P. T. M., Dolfing, J., Eshuis, E. J., and Scholten-Koerselman, I. J.: 1979, Microbial. Ecol. 5, 139.