Aquatic Toxicology 48 (2000) 275–289

pH-dependent toxicity of copper and uranium to a tropicalfreshwater alga (Chlorella sp.)

Natasha M. Franklin a,b,*, Jennifer L. Stauber a, Scott J. Markich c,Richard P. Lim b

a Centre for Ad6anced Analytical Chemistry, CSIRO Energy Technology, PMB 7, Bangor, NSW 2234, Australiab Department of En6ironmental Sciences, Uni6ersity of Technology, Sydney, PO Box 123, Broadway, NSW 2007, Australia

c En6ironment Di6ision, Australian Nuclear Science and Technology Organisation, PMB 1, Menai, NSW 2234, Australia

Received 25 November 1998; received in revised form 20 April 1999; accepted 11 May 1999

Abstract

Copper (Cu) and uranium (U) are of potential ecotoxicological concern to tropical freshwater organisms innorthern Australia as a result of mining activity. No local data on the toxicity of these metals to tropical freshwateralgae are currently available. The aim of this study was to investigate the effect of pH (5.7 and 6.5) on the toxicityof Cu and U to the green alga Chlorella sp. in a synthetic softwater representative of fresh surface waters insandy-streams of tropical northern Australia. The effects of Cu and U on algal growth (cell division) rate after a 72-hexposure were determined. Intracellular and extracellular (membrane-bound) metal concentrations at the two selectedpH values were also compared. Based on the 72-h minimum detectable effect concentrations (MDEC), Chlorella sp.was approximately 20-fold more sensitive to Cu (0.7 and 1.4 mg l−1 at pH 6.5 and 5.7, respectively) than U (13 and34 mg l−1 at pH 6.5 and 5.7, respectively), and more sensitive than other Australian tropical freshwater organisms.The toxicity of Cu and U was highly pH-dependent. Copper concentrations required to inhibit growth (cell division)rate by 50% (72-h EC50) increased from 1.5 to 35 mg l−1 as the pH decreased from 6.5 to 5.7. Similarly, the 72-h EC50

values for U increased from 44 to 78 mg l−1 over the same pH range. Calculation of Cu and U speciation using thegeochemical model HARPHRQ, showed that differences in the concentrations of the free metal ions (Cu2+ andUO2

2+) were only minimal (B10%) between pH 5.7 and 6.5. The decreased toxicity at pH 5.7 was due to lowerconcentrations of cell-bound and intracellular Cu and U compared to those at pH 6.5. These results are explained interms of the possible mechanism of competition between H+ and the metal ion at the cell surface. © 2000 ElsevierScience B.V. All rights reserved.

Keywords: Toxicity; Algae; Copper; Uranium; pH

www.elsevier.com/locate/aquatox

* Corresponding author. Tel.: +61-2-9710-6833; fax: +61-2-9710-6837.

E-mail address: natasha.franklin@det.csiro.au (N.M.Franklin)

1. Introduction

Metal contamination of aquatic environmentsoccurs as a result of human activities and affects

0166-445X/00/$ - see front matter © 2000 Elsevier Science B.V. All rights reserved.

PII: S 0166 -445X(99 )00042 -9

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organisms at the biochemical, cellular, populationand community level. In tropical Australian fresh-waters, copper (Cu) and uranium (U) inputs frommining activities are of potential ecotoxicologicalconcern (Markich and Camilleri, 1997). Althoughavailable local data on the toxicity of these metalsto animals are gradually increasing, there is littleinformation on the toxicity of Cu and U to tropi-cal freshwater algae. Unicellular algae are an im-portant component of aquatic systems as they arethe foundation of most food chains and accountfor much of the production base of freshwaterand marine ecosystems. As primary producers,algae fix a major portion of the Earth’s carbonand generate, via photosynthesis, much of theoxygen in our atmosphere. Any adverse impact onalgae is likely to affect organisms at highertrophic levels and may have important conse-quences for the health of the whole aquaticecosystem.

It is now well established that the physico-chemical form of a metal (speciation) is a criticalfactor controlling metal bioavailability, andhence, toxicity in natural waters. Generally, toxic-ity correlates with the concentration of the freemetal ion (e.g. Cu2+) (Campbell, 1995); however,some exceptions to this free-ion activity modelhave been reported, including lipid-soluble coppercomplexes (Florence and Stauber, 1986; Camp-bell, 1995).

Copper has been found to be one of the moretoxic metals to unicellular algae. For example,numerous comparative studies have reported thetoxicity of Cu to a variety of algal species to begreater than that of cadmium (Cd) (Jayaraj et al.,1992; Guanzon et al., 1994), lead (Pb) (Kessler,1986; Starodub et al., 1987), chromium (Cr)(Wong and Chang, 1991), nickel (Ni) (Wong andChang, 1991; Jayaraj et al., 1992) and thallium(Tl) (Canterford and Canterford, 1980). At con-centrations above those required for optimalgrowth, Cu has been shown to inhibit growth(Aliotta et al., 1983; Baker et al., 1983; Kessler,1986), extend the lag phase of cultures (SteemannNielsen and Wium-Anderson, 1970; Bartlett et al.,1974) and interfere with several cellular processesincluding photosynthesis, respiration, enzyme ac-tivity, pigment synthesis and cell division (Stauber

and Florence, 1987; Ahmed and Abdel-Basset,1992; Guanzon et al., 1994). Unfortunately, manyof these studies have been carried out in highnutrient culture medium which contains chelatorsand adsorbants that can reduce metal toxicity. Toovercome this problem, recent protocols for toxic-ity testing with algae recommend the use of mini-mal media enriched only with nitrate andphosphate (Stauber, 1995).

The speciation and toxicity of U in naturalwater, like Cu, is controlled by a variety of fac-tors, including pH, and complexation by organicand inorganic ligands. To our knowledge, thereare no data on the toxicity of U to unicellularalgae. The toxicity of U to a number of tropicalfreshwater invertebrates and fish has been studied,and EC50 values range from 40 to 1300 mg l−1

(Markich and Camilleri, 1997).Water pH is an important factor affecting the

toxicity of metals to algae, although a relationshipbetween the two has been difficult to establish.Some authors have shown an increase in metaltoxicity with decreasing pH (Starodub et al., 1987;Rai et al., 1993, 1994) due to the predominance ofthe free metal ion (Baker et al., 1983; Borgmann,1983; Starodub et al., 1987). Conversely, otherstudies have shown that the toxicity of metals [e.g.Cd, Cu, and zinc (Zn)] decreases markedly withdecreasing pH (Steemann Nielsen and Kamp-Nielsen, 1970; Peterson et al., 1984; Campbell andStokes, 1985), the consequence of reduced metaluptake due to competition with H+ at the cellmembrane surface. The objective of the presentstudy was to determine the effect of pH (5.7 and6.5) on the toxicity of Cu or U to a tropicalfreshwater green algae, Chlorella sp. Intracellularand extracellular (membrane-bound) metal con-centrations at the two selected pH values werealso compared.

2. Materials and methods

2.1. Algal stock cultures

The unicellular freshwater green alga Chlorellasp. (Chlorophyceae) was isolated from surfacewater collected within Kakadu National Park

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 277

(13°S, 132°E), Alligator Rivers Region in theNorthern Territory of Australia (Padovan, 1992).This isolate could not be identified, and is possi-bly a new Chlorella species (H. Ling, pers. com-mun.). An axenic culture was established in MBLmedium (Stein, 1973) at the CSIRO Division ofEnergy Technology (Sydney, Australia). Thestock cultures were incubated at 2791°C on a12-h light, 12-h dark cycle (Philips TL 40 W coolwhite fluorescent lighting, 75 mmol photons m−2

per s). Cells in exponential phase growth wereused for all experiments.

2.2. Growth inhibition bioassays

The toxicity of Cu or U to Chlorella sp. wasdetermined using a recently developed 72-hgrowth inhibition bioassay (Franklin et al., 1998)(Table 1). The test medium used was a syntheticsoftwater (Table 2) having a chemical composi-

Table 2Measured concentrations of inorganic components of syntheticMagela Creek water (n=4–9)

Mean value9S.D.Parameter

Na (mg l−1) 1.0090.120.4690.02K (mg l−1)

Ca (mg l−1) 0.4790.01Mg (mg l−1) 0.6790.02Cl (mg l−1) 1.3590.03

2.8590.06SO4 (mg l−1)Fe (mg l−1) B50.0Al (mg l−1) 34.092.42

10.090.06Mn (mg l−1)B1.00U (mg l−1)

0.6090.07Cu (mg l−1)

tion similar to that of sandy, braided-streams intropical northern Australia during the wet season(November–April). Water quality characteristicswere specifically based on the chemical composi-tion of surfaces waters of Magela Creek (AlligatorRivers Region, Northern Territory) (Markich,1998). These waters are very soft (2–4 mg l−1 asCaCO3) and slightly acidic (pH 6.0), with verylow concentrations of cations and anions (Hart etal., 1987).

The test containers were 250-ml Erlenmeyerflasks coated with a silanizing solution (Coatasil,Ajax), and acid-washed (10% concentratedHNO3) before use. To each flask, 0.5 ml of 26mM sodium nitrate (15 mg NO3 l−1) and 0.05 mlof 1.3 mM potassium dihydrogen phosphate (0.15mg PO4 l−1) were added to 54.5 ml of filteredsynthetic softwater.

Copper stock solutions (5 and 100 mg l−1) wereprepared from copper sulfate (CuSO4·5H2O) andacidified (pHB2) by the addition of HCl. Copperconcentrations tested ranged from 1.25 to 640 mgl−1. For the U bioassays, a 10 mg l−1 stocksolution was prepared from uranyl sulfate(UO2SO4·3H2O), with test concentrations rangingfrom 5 to 320 mg l−1. Controls, together with atleast five metal concentrations (each in triplicate)were prepared. Sub-samples (5 ml) were takenfrom each flask at the beginning of each toxicitytest and acidified (pHB2), prior to determinationof total dissolved Cu by graphite furnace atomicabsorption spectrometry (GFAAS, Perkin Elmer

Table 1Summary of standard bioassay protocol for the tropicalChlorella sp. growth inhibition bioassay

1. Test type Static2. Temperature 2791°C

Cool white fluorescent3. Light qualitylighting115–125 mmol photons PAR4. Light intensitym−2 per s

5. Photoperiod 12-h light, 12-h dark cycle6. Test chamber size 250 ml

50 ml7. Test solution volume8. Renewal of test solutions None9. Age of test organisms 4–5 days

2–4×104 cells ml−110. Initial cell density in testchambers

311. No. of replicatechambers/concentration

12. Shaking rate Twice daily by handSynthetic softwater+15 mg13. Test mediuml−1 NO3+0.15 mg l−1 PO4

14. Metal concentrations Minimum of five and acontrol

15. Test duration 72 h16. Test endpoint Growth (cell division) rate

Control growth rate17. Test acceptability1.490.3 (pH 5.7) and1.790.2 (pH 6.5) doublingsday−1; C.V. in controlsB20%

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Model 4100ZL) and U by inductively coupledplasma atomic emission spectrometry (ICPAES,Labtam Model 8410). Measured rather thannominal concentrations were used to calculatetoxicity endpoints.

Each flask was inoculated with 2–4×104 cellsml−1 of prewashed Chlorella suspension. Flaskswere incubated at 2791°C on a 12-h light, 12-hdark cycle (Philips TL 40 W cool white fluores-cent lighting, 120 mmol photons PAR m−2 pers). To minimise the problem of increasing pHover the test period, as a result of the utilisationof carbon dioxide and bicarbonate by algal cells,the pH of the test solution was adjusted twicedaily by the addition of dilute HCl. Algal celldensity was determined daily using a CoulterMultisizer IIE Particle Analyser with a 70-mmaperture with correction for background particlecount.

2.3. The effect of pH on Cu and U toxicity

To determine an appropriate pH at which totest the toxicity of Cu and U, control growthrates of Chlorella sp. were compared over arange of pH values (pH 5.0–7.0) typical of Aus-tralian tropical freshwaters. A series of algalgrowth inhibition bioassays were conducted atpH 5.7 and 6.5 to determine the effect of H+

concentration on metal toxicity. pH values of5.0 and 7.0 were considered inappropriate fortesting due to high variability in control growthrates (coefficient of variation\20%) or evidenceof a lag phase. To ensure that the growth rateof Chlorella sp. exposed to Cu and U was re-producible, a range-finding test was performedfollowed by three definitive toxicity tests at bothpH 5.7 and 6.5.

2.4. Intracellular and extracellular metaldetermination

To investigate the mechanism by which pHmay influence metal toxicity, intracellular andextracellular Cu and U concentrations were de-termined after a 72-h exposure. Quadruplicateflasks were set up at both pH 5.7 and 6.5 fol-lowing the standard test protocol (Franklin et

al., 1998). Three Cu treatments were prepared ateach pH (i.e. pH 5.7: 10, 80 and 640 mg l−1;pH 6.5: 10, 80 and 160 mg l−1) based on thefinal concentration-response curves. Uraniumconcentrations used were 40, 120 and 250 mgl−1 at pH 5.7 and 6.5. One flask from eachtreatment was used to determine the daily celldensity for 3 days. This flask acted as a testcontrol to allow a comparison with earlierbioassays. The remaining three flasks per treat-ment were counted only on day 3 to allow suffi-cient volume for intracellular and extracellularCu or U determination. All flasks were pH-ad-justed twice daily with 0.01 M HCl to maintainthe pH within 0.5 pH unit.

At the completion of the bioassay (day 3), theremaining test medium was retained. In a Class-100 clean room, 40.090.1 g of solution fromeach flask was decanted into 50-ml Teflon cen-trifuge tubes which had been presoakedovernight in 50% concentrated HNO3 and thor-oughly washed with Milli-Q water. Sampleswere centrifuged for 20 min at 3500 rev./minand 20°C in a Joun CR4.11 refrigerated cen-trifuge. The supernatant solutions (20 ml) werepipetted into clean, acid-washed (10% concen-trated HNO3) polycarbonate vials and acidifiedwith 40 ml of concentrated HNO3 (MerckSuprapur). These samples were analysed for dis-solved Cu by GFAAS and dissolved U by IC-PAES. A further 10 ml of the supernatant wasdiscarded and the remaining 10 ml re-cen-trifuged for 5 min. This additional centrifuga-tion step was necessary so that the cells formeda distinct pellet. The supernatant solution wasdiscarded and the algal pellet was re-suspendedin 20 ml of 0.02 M ethylenediaminetetraaceticacid di-sodium salt (EDTA) and shaken for ap-proximately 30 s which removed the Cu or Ubound to the outside of the cells (Florence andStauber, 1986). Preliminary experiments showedthat this EDTA treatment did not cause cell ly-sis. The samples were then centrifuged for a fur-ther 20 min at 4000 rev./min and thesupernatant removed and retained for Cu or Uanalysis. No acidification was necessary. Thesesamples were referred to as the ‘extracellular Cuor U’ fraction.

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 279

The algal pellets were left to dry in a laminar flowhood for 2 days and then acid-digested with 2 mlof concentrated HNO3. After standing for 30 min,samples were microwaved on a power setting of 1(90 W) for 5 min. After cooling to room tempera-ture, the samples were made up to 20 ml withMilli-Q water and analysed for Cu by GFAAS andU by ICPAES. These samples were referred to asthe ‘intracellular Cu or U’ fraction.

To calculate the mass balance of metals, any Cuor U that may have adsorbed to the walls of theglass flask over the duration of the toxicity test wasdetermined by emptying flasks and filling with 50ml of 0.2% HNO3 for 16 h. Copper in this acid wasanalysed by GFAAS and U by ICPAES, and werereferred to as the ‘flask-bound Cu or U’ fraction.

2.5. Statistical analysis

Linear regression analysis was used to fit log10

cell density vs time (h) for each sample. The growth(cell division) rate was determined from the regres-sion slope. The toxicity of Cu and U was expressedas the effective concentration giving 50% reduction(EC50) in algal growth rate over 72 h compared tothe controls. This was calculated using theTrimmed Spearman–Karber analysis (ToxCalcVersion 5.0.14, Tidepool Software). After testingthe data for normality and homogeneity of vari-ances, Dunnett’s Multiple Comparison Test wasused to determine which treatments were signifi-cantly different from the controls. Alternativestatistical measures to the no observable effectconcentration (NOEC) and lowest observable ef-fect concentration (LOEC) were calculated. The10% bounded effect concentration (BEC10), analternative to the NOEC, was estimated using theapproach described by Hoekstra and van Ewijk(1993). The minimum detectable effect concentra-tion (MDEC), an alternative to the LOEC, wasestimated using the approach described by Ahsan-ullah and Williams (1991). Tests for significancebetween treatments (e.g. comparison of intracellu-lar and extracellular Cu and U concentrations)were determined using a one-way analysis of vari-ance (ANOVA) or Student’s t-test. Significancelevels were tested at the P=0.05 level, unlessotherwise indicated.

Table 3Comparison of test endpoints for Chlorella sp. exposed to Cuat pH 5.7 and 6.5

Test endpoint Cu (mg l−1)

pH 5.7 pH 6.5

1.790.2Mean growth rate 1.490.3(doublings day−1)9S.D.

72-h EC50 (95% C.I.) 1.5 (0.8–2.8)35 (28–42)1.1BEC10 0.6

MDEC 1.4 0.7

3. Results

3.1. Comparison of control algal growth rates atpH 5.0, 5.7, 6.5 and 7.0

Chlorella sp. grew well at all pH values tested,with \1.0 doubling day−1. A lag phase of approx-imately 24 h was observed at pH 5.0 and 7.0 beforethe onset of growth. At pH 5.0, the % C.V. was\20%. A one-way ANOVA showed that cells grewsignificantly (P50.05) better at pH 6.5 than at pH5.7 (Tables 3 and 4). No significant (P\0.05)difference in the algal growth rate was detectedbetween pH 5.0 and 5.7 or between pH 6.5 and 7.0.Due to the high variability found at pH 5.0, andthe apparent lag phase observed at pH 5.0 and pH7.0, pH values of 5.7 and 6.5 were chosen toinvestigate pH effects on metal toxicity. pH controlin the unbuffered water was reasonable (B0.5 pHunit increase) for all pH treatments from days 0–2of the test; however, an increase of \0.5 pH unitwas typically observed in the controls by day 3 asa result of high algal cell densities.

Table 4Comparison of test endpoints for Chlorella sp. exposed to U atpH 5.7 and 6.5

Test endpoint U (mg l−1)

pH 6.5pH 5.7

Mean growth rate 1.490.3 1.790.2(doublings day−1)9S.D.

72-h EC50 (95% C.I.) 44 (39–49)78 (71–83)11BEC10 2113MDEC 34

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289280

3.2. Copper toxicity

The effect of increasing Cu concentrations onthe growth rate of Chlorella sp. at pH 5.7 and 6.5is shown in Fig. 1a and Fig. 1b, respectively. Inthe absence of Cu, growth rates of 1.4 doublingsday−1 (pH 5.7) and 1.7 doublings day−1 (pH 6.5)were obtained, with a % C.V.B20%, indicatingtest acceptability. Growth of the alga at each pHtypically decreased with increasing Cu concentra-tion. At pH 5.7, low concentrations of Cu (1.4 mgl−1) caused only minor growth rate inhibition(Fig. 1a). As Cu concentrations increased (2.4–19mg l−1), a gradual decrease in algal growth ratewas observed, reducing growth rates to approxi-mately 55% of the control. Further increases inCu concentration up to a threshold level of 160 mgl−1 (Fig. 1a) had little additional effect (B10%)on algal growth rate. Beyond these Cu concentra-tions, pronounced inhibition occurred until zerogrowth rate at 600 mg l−1. The calculated 72-hEC50 value was 35 mg l−1 Cu (28–42 mg l−1; 95%C.I.).

The concentration-response curve for Chlorellasp. is not the typical sigmoidal curve shown bymany algal species in the presence of a metaltoxicant. The BEC10 and MDEC values were 1.1and 1.4 mg l−1, respectively. Control of pH in theunbuffered water was reasonable throughout thebioassays, typically increasing B0.5 pH unit over72 h. For all Cu toxicity tests, measured Cuconcentrations at the beginning of the test were inclose agreement with nominal concentrations (B10% difference).

Chlorella sp. was extremely sensitive to Cu atpH 6.5 (Table 3). The concentration-responsecurve was initially very steep with pronouncedgrowth inhibition at low Cu concentrations (Fig.1b). The BEC10 and MDEC values were 0.6 and0.7 mg l−1, respectively. These values were similarto background levels of Cu (0.6 mg l−1) in thesynthetic water (Table 2). The concentration-re-sponse curve was relatively flat over Cu concen-trations from 1.2 to 80 mg l−1. The 72-h EC50

value was 1.5 mg l−1 (0.8–2.8 mg l−1; 95% C.I.).The pH was maintained within B0.5 pH unitover 72 h for all Cu treatments. However, the pH

Fig. 1. Growth rate inhibition of Chlorella sp. exposed to Cufor 72 h at (A) pH 5.7 and (B) pH 6.5. Data points representthe mean9S.D. of three tests (except for 1.4, 306 and 611 mgl−1 at pH 5.7, and 1.2 and 160 mg l−1 at pH 6.5, whichrepresent only one test).

in controls increased by about 1 pH unit by theafternoon of day 2 of the test.

It is clear from the results that Cu was signifi-cantly (P50.05) more toxic to Chlorella sp. at pH6.5 than at pH 5.7. Increasing the pH of the test

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 281

Table 5Surface-bound and intracellular Cu for Chlorella sp. at pH 5.7 and 6.5a

Measured Cu (mg l−1) Intracellular Cu Extra/Intra CuAdded Cu Extracellular Cu(mg l−1) (×10−6 ng cell−1)(×10−6 ng cell−1)

pH 5.7B0.5 B0.5B0.5 –0

9.310 5.1390.26 3.5490.63 0.6990.1318.190.76 29.0911.074 1.5990.5780

599640 121933.7 3529314 2.6191.66

pH 6.5B0.5 B0.5B0.5 –0

9.410 10.992.05 5.4490.30 0.5690.031.894.57 56.1911.180 1.9090.708469.4912.7 10396.69165 1.5490.34160

a Values are reported as the mean9S.D.

medium resulted in a substantial reduction in the72-h EC50 value from 35 to 1.5 mg l−1 (Table 3).To explain this 20-fold increase in Cu toxicity atpH 6.5, the cellular Cu content of the algae wasdetermined at the completion of the toxicity testat each pH value.

Uptake of Cu by Chlorella sp. at pH 5.7 and6.5 is summarised in Table 5. The cellular Cucontent (i.e. intracellular+extracellular Cu) ofChlorella sp. increased with increasing Cu concen-tration in the test media at both pH 5.7 and 6.5.On a per cell basis, the ratio of extracellular/intra-cellular Cu was similar at both pH values. At 10mg l−1, up to 60% of the total cellular Cu waslocated intracellularly, and 40% was bound to thecell surface. In contrast, at higher Cu concentra-tions (e.g. 640 mg l−1) the majority of cellular Cu(75%) was bound to the cell surface with only25% located intracellularly. Differences in the ra-tio of extracellular/intracellular Cu were not sig-nificant (P\0.05) among Cu treatments.

A direct comparison of the cellular Cu contentat pH 5.7 and 6.5 could be made for the 10 and 80mg l−1 treatments (Table 5). Extracellular Cu wassignificantly (P50.05) higher at pH 6.5 than at5.7, with approximately 2.5 times more Cu boundto the cell surface at pH 6.5. Similarly, intracellu-lar Cu was significantly (P50.05) higher (two-fold) at pH 6.5 than at 5.7. Toxicity (i.e. growthrate inhibition) correlated with both intracellularand extracellular Cu, i.e. the more Cu bound at

the cell surface, the more Cu penetrated the celland the greater the toxicity.

A mass balance of the Cu added to each flaskwas also calculated, with recovered Cu\90%. Atthe higher Cu concentrations (80, 160 and 640 mgl−1) most Cu at the end of the bioassay was stillpresent in solution (65–90% dissolved Cu). Theintracellular and extracellular Cu fractions eachrepresented less than 10% of the total Cu added.A considerable portion of Cu (up to 25%) ad-sorbed to the walls of the flasks throughout thetest despite silanization of the glass flasks prior tothe bioassay.

3.3. Uranium toxicity

Growth of Chlorella sp. typically decreasedwith increasing U concentrations at pH 5.7 and6.5, respectively (Fig. 2a and Fig. 2b). In theabsence of U, growth rates of 1.4 doublingsday−1 (pH 5.7) and 1.7 doublings day−1 (pH 6.5)were obtained, with the % C.V.B20%, indicatingtest acceptability (Table 4). At pH 5.7, U concen-trations of up to 34 mg l−1 had no significant(P\0.05) effect on growth rate. The concentra-tion-response curve was particularly steep fromabout 60 mg l−1, with relatively small changes inU concentrations causing pronounced growth in-hibition (Fig. 2a). Unlike the results obtained inthe presence of Cu, Chlorella sp. showed a typicalsigmoidal response in the presence of increasing U

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289282

concentrations. The 72-h EC50 value at pH 5.7was 78 mg l−1 (71–83 mg l−1; 95% C.I.). TheBEC10 and MDEC values were 21 and 34 mg l−1,respectively. Control of pH in the unbufferedwater was reasonable over 48 h, rising B0.5 pH

unit. Measured U concentrations were 10–20%lower than nominal concentrations, indicating aloss of U to the glass walls of the flasks, despitesilanization prior to the test.

Chlorella sp. showed increased sensitivity to Uat pH 6.5 relative to pH 5.7 (Fig. 2b). At concen-trations\8 mg l−1, small increases in U causedpronounced growth inhibition. The 72-h EC50

value was 44 mg l−1 (39–49 mg l−1; 95% C.I.).The BEC10 and MDEC values were 11 and 13 mgl−1, respectively. Throughout the test, the pH wasmaintained within 0.5 units of pH 6.5, with theexception of the controls in which pH increased to7.2 by day 3 of the test.

Uptake of U by Chlorella sp. at pH 5.7 and 6.5is summarised in Table 6. The cellular U content(i.e. intracellular+extracellular U) increased withincreasing U concentration at both pH 5.7 and6.5. For example, increasing the U concentrationfrom 40 to 250 mg l−1 resulted in an eightfoldincrease in both intracellular and extracellular U(pH 5.7 and 6.5). On a per cell basis, the ratio ofextracellular/intracellular U was not significantly(P\0.05) different between pH 5.7 and 6.5 (at 40and 120 mg l−1 U), with less than 30% of the totalcellular U located intracellularly. However, at 250mg l−1 U, significantly (P50.05) more U waslocated extracellularly relative to intracellularconcentrations at pH 6.5 than at pH 5.7, suggest-ing possible changes in membrane permeability.

At all U concentrations tested, intracellular Uwas significantly (P50.05) greater at pH 6.5 thanat pH 5.7, with approximately two times more Ulocated inside the cell. The concentration of sur-face-bound U was also greater at pH 6.5 com-pared to pH 5.7, however this difference was onlysignificant for the 250 mg l−1 U treatment as aresult of the large variation between replicates inthe other treatments.

A mass balance of the U added to the testflasks was also calculated, with recovered U\90%. At the higher U concentrations (120 and 250mg l−1) most U at the end of the bioassay was stillpresent in solution (50–70% dissolved U). Theintracellular and extracellular U fractions eachrepresented less than 10% of the total U added.Like Cu, a significant fraction of U (up to 40%)adsorbed to the walls of the flasks throughout thetest.

Fig. 2. Growth rate inhibition of Chlorella sp. exposed to Ufor 72 h at (A) pH 5.7 and (B) pH 6.5. Data points representthe mean9S.D. of three tests (except for 8.0, 33, 217 and 265mg l−1 at pH 5.7, and 4.0, 205 and 271 mg l−1 at pH 6.5,which represent only one test).

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 283

Table 6Surface-bound and intracellular U for Chlorella sp. at pH 5.7 and 6.5a

Intracellular UMeasured UAdded U Extracellular U Extra/Intra U(mg l−1) (×10−6 ng cell−1)(×10−6 ng cell−1)(mg l−1)

pH 5.70 B0.5 B0.5 B0.5 –

3640 16.893.60 21.7910.7 1.2590.4069.695.20 116939.0110 1.6590.47120

230250 13596.40 170936.8 1.3290.21

pH 6.5B0.5 B0.5B0.5 –0

3340 31.193.80 70.5926.1 2.2790.8790.993.00120 196939.798 2.1790.51247920.6 5439121211 2.2090.49250

a Values are reported as the mean9S.D.

Cells exposed to U had a significantly (P50.05) larger mean cell diameter compared to con-trol cells, and this increase in cell surface area wassimilar at both pH values. Since U-treated cellswere larger than control cells, intracellular andextracellular U was also expressed on a cell sur-face area basis (data not shown). At 120 mg l−1,significantly (P50.05) more intracellular and ex-tracellular U was detected at pH 6.5 compared topH 5.7. This difference was not detected for the40 mg l−1 treatment on a cell surface area basisdue to the large variability in the analyses at thisU concentration.

4. Discussion

The tropical Australian freshwater algaChlorella sp. was highly sensitive to Cu and U ina synthetic softwater typical of the region, makingit a suitable standard test organism for site-spe-cific testing of mine wastewaters.

It was difficult to compare the toxicity of Cu toChlorella sp. with other freshwater algae becausemost other bioassays have been carried out inculture media containing chemicals which com-plex Cu, thereby underestimating its toxicity.However, several comparable studies were avail-able. Stauber and Florence (1989) reported a 72-hEC50 value of 16 mg l−1 for the temperate fresh-water alga Chlorella protothecoides in synthetic

softwater (pH 7.5). In high nutrient MBLmedium, the same authors reported a 72-h EC50

value of \200 mg l−1, emphasising the impor-tance of the test medium in deriving toxicity data.For the tropical freshwater isolate Chlorella sp. 12from Papua New Guinea, a 72-h EC50 value of8.4 mg l−1 was obtained in synthetic softwater(Franklin et al., 1998). It should be noted that allof these studies were performed at pH 7.5, andtherefore, are not directly comparable to thisstudy with Chlorella sp. because Cu toxicity ishighly pH-dependent (Table 3). However, giventhe constraints in making such a comparison, thepresent study shows that Chlorella sp. is at leastas sensitive as other freshwater algae to Cu.

The relative sensitivity of Chlorella sp. to Cuwas compared with other tropical freshwaterbiota (Table 7). For comparative purposes, onlytoxicity data that were derived under similar envi-ronmental conditions (i.e. softwater, slightlyacidic, low alkalinity and conductivity) are used.

The relative sensitivity of these various testspecies to Cu is best compared at pH 5.7–6.0,over which range the speciation of Cu, as pre-dicted using the geochemical speciation modelHARPHRQ, is similar. Based on EC50 values,Chlorella sp. was less sensitive to Cu than thegreen hydra, the bivalve and the gudgeon fish.However, Chlorella sp. was inhibited at concen-trations as low, or lower, than the other tropicalanimals mentioned [i.e. MDEC values (mg l−1

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289284

Table 7Comparative toxicity of Cu to Australian tropical freshwater biota

pH Cu toxicity (mg l−1)Species ReferenceEndpoint

5.7Microlgae (Chlorella sp.) 359772-h EC50 This study6.5 1.5916.0Cnidaria (Hydra 6iridissima) 4.090.396-h EC50 Markich and Camilleri (1997)6.0 239196-h LC50 Markich and Camilleri (1997)Chordata (Mogurnda mogurnda)

Mollusca (Velesunio angasi ) 5.896-h EC50 8.990.5 Markich (1998)

Cu): 1.4 (alga, pH 5.7), 1.6 (hydra, pH 6.0), 6.7(bivalve, pH 5.8), 7.0 (fish, pH 6.0)]. Furthermore,Chlorella sp. was extremely sensitive to Cu at pH6.5 (72-h EC50=1.5 mg l−1). Further work isneeded to address the influence of pH on thetoxicity of Cu to other tropical freshwater biota.

This study provides the first known data on thetoxicity of U to a freshwater alga. The 72-h EC50

value was 78 mg l−1 at pH 5.7 and 44 mg l−1 atpH 6.5 (Table 4). Although there are no compara-ble toxicity data for unicellular algae, there aresome data for Australian tropical freshwater ani-mals (Table 8). Despite the different test durationsused, the relative order of sensitivity is: Chlorellasp.]water flea\green hydra\bivalve\fish.

Based on EC50 values, the toxicity of Cu and Uwas compared on a molar basis, to account forthe difference in molecular mass. Uranium wasnearly two times more toxic than Cu at pH 5.7.At pH 6.5, Cu was the more toxic (×9) of thetwo metals. Other studies (Poston et al., 1984;Markich and Camilleri, 1997) have shown that atpH 5.7, Cu is more toxic than U.

It is evident from the concentration-responsecurves obtained for U and Cu that Chlorella sp.responds differently to the two metals. In thepresence of U, Chlorella sp. showed a sigmoidal-type response, resulting in a sharp decline in algalgrowth rate above a certain threshold metal con-centration. This response is typical of mostaquatic organisms in the presence of a toxicant(Rand et al., 1995). However, the concentration-response curve for Cu was particularly flat over awide metal concentration range (Fig. 1) until athreshold level, indicating that increasing the Cuconcentration had little additional effect on algalgrowth rate. It is possible that this flat area of the

concentration-response curve may be due todetoxification of Cu by the algal cells. Certainalgal species, including Scenedesmus actutiformisand Chlorella fusca, have been shown to produceorganic substances that reduce the bioavailableCu concentration if they are released extracellu-larly in sufficient amounts (Robinson, 1989). Cellwall exclusion of Cu and intracellular detoxifica-tion by binding to phytochelatins or polyphos-phate has also been demonstrated in a variety offreshwater algal species in response to elevated Cuconcentrations (Twiss et al., 1993; Knauer et al.,1997). This mechanism of detoxification could beinvestigated further by comparing intracellularand extracellular Cu levels at Cu concentrationsalong the flat portion of the concentration-re-sponse curve.

The toxicity of Cu or U to freshwater biota ismarkedly affected by key water quality variables,such as pH, water hardness and dissolved organiccarbon concentration. In natural waters, thesefactors modify toxicity by either increasing ordecreasing the concentration required to producea particular biological response (Chapman, 1995).Based on EC50 values, an 18-fold increase in Cutoxicity and a twofold increase in U toxicity wasobserved by increasing the pH from 5.7 to 6.5,supporting previous observations in the literature.Peterson et al. (1984) found that total Cu toxicityto the unicellular green alga Scenedesmus quadri-cauda, increased almost 76-fold (concentration re-quired to give 50% inhibition of phosphate uptakedecreased from 7.6 to 0.1 mmol l−1) with anincrease in pH from 5.0 to 6.5. Similarly, Macfieet al. (1994) found that the toxicity of Cu (basedon growth rate inhibition) to Chlamydomonasreinhardtii, increased sevenfold at pH 7 comparedto pH 5.

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 285

Table 8Comparative toxicity of U to Australian tropical freshwater biota

pH U toxicity (mg l−1)Species ReferenceEndpoint

5.7Microlgae (Chlorella sp.) 789672-h EC50 This study6.5 44956.5Crustacea (Moinodaphnia macleayi ) 3993120-h EC50 Markich and Camilleri (1997)6.0 959596-h EC50 Markich and Camilleri (1997)Cnidaria (Hydra 6iridissima)

96-h LC50Chordata (Mogurnda mogurnda) 6.0 1360950 Markich and Camilleri (1997)5.8 254915Mollusca (Velesunio angasi ) Markich (1998)96-h EC50

To interpret metal toxicity results at differentpH values it is important to know how Cu specia-tion in solution changes with pH (Peterson et al.,1984). The speciation of Cu in the test water atpH 5.7 and 6.5 was predicted using HARPHRQ.The input parameters for HARPHRQ were basedon physico-chemical data measured in the testwater. Equilibrium constants were taken fromMarkich (1998). Measured concentrations of inor-ganic components of the synthetic test water typi-cally varied by B10% from the nominalconcentrations. At pH 5.7, Cu was predicted tooccur predominantly (98%) as the free hydratedmetal ion (Cu2+) with only a very small contribu-tion (1.6%) from CuOH+.

It is generally accepted that metal toxicity isgoverned by the activity of the free hydratedmetal ion, rather than total metal concentration[i.e. the free-ion activity model (FIAM); Camp-bell, 1995). One exception to the FIAM is lipid-soluble metal complexes; however, suchcomplexes typically represent B1% of the totaldissolved metal concentration (Phinney and Bru-land, 1994). Also, the rate at which a metal com-plex dissociates at the solution/cell membraneinterface can influence toxicity (Florence andStauber, 1991). Therefore, at pH 5.7, Cu is poten-tially in its most bioavailable, and hence toxic,form to Chlorella sp. Increasing the pH of the testmedium to 6.5 reduced the predicted percentageof Cu2+ to 89%, while the proportion of CuOH+

increased to 9.1%. Since Cu2+ is highly correlatedwith toxicity, then according to the FIAM, Cushould be slightly more toxic to Chlorella sp. atpH 5.7 than at pH 6.5. However, the results fromthe growth inhibition tests showed the opposite;i.e. that Cu was more toxic at pH 6.5.

There is growing evidence to suggest that hy-drogen ions may decrease the toxicity of metals bycompetitively excluding them from binding tocell-surface ligands (Peterson et al., 1984; Camp-bell and Stokes, 1985; Parent and Campbell,1994). For example, Pawlik et al. (1993) found Cdinhibition of photosynthesis to the cyanobac-terium Synechocystis aquatilis greater at pH 7.0than at pH 5.5, corresponding to increased Cduptake. The present study supports the hypothesisthat H+ may competitively exclude Cu from bind-ing at the cell surface, as less Cu was bound to thecell at higher H+ concentrations (i.e. lower pH).The intracellular Cu content of Chlorella sp. wasalso significantly (P50.05) lower at pH 5.7 thanat pH 6.5. As the pH increases, more Cu binds tothe cell surface and a twofold increase in intracel-lular Cu was measured. The toxicity results confi-rmed that increased Cu toxicity to Chlorella sp. atpH 6.5 was due to increased metal uptake into thecell.

The speciation of U in the test water at pH 5.7and 6.5 was also predicted using HARPHRQ. Incontrast to the relatively simple speciation pre-dicted for Cu at pH 5.7 and 6.5, the speciation ofU in the test water is complex. At pH 5.7, the freeuranyl ion (UO2

2+) constitutes a minor propor-tion of the total U concentration (i.e. 12% at 0.1mg l−1 declining to 1.3% at 1000 mg l−1). Incontrast, the proportion of polymeric uranyl spe-cies, including (UO2)3(OH)5

+, (UO2)3(OH)7− and

(UO2)2(OH)3CO3− increase as the total U concen-

tration increases. These increases are compensatedby a reduction in the relative proportions ofmonomeric uranyl species [i.e. UO2(OH)+ andUO2CO3]. Increasing the pH of the test mediumto 6.5 increased the formation of polymeric uranyl

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289286

species. The proportion of free uranyl ion (UO22+)

was found to be insignificant (B1%) at pH 6.5.Markich et al. (1996) showed that the toxicity ofU to a tropical freshwater bivalve (V. angasi ) wasgoverned by UO2

2+ and UO2OH+, rather than thesum of the inorganic uranyl species.

It is apparent from the 72-h EC50 values ob-tained in the present study, that the toxicity of Uto Chlorella sp. is not solely dependent on thespeciation of U in the test waters over the pHrange 5.7–6.5. If this were true, then U toxicitywas expected to be highest at pH 5.7 compared topH 6.5, corresponding to a greater proportion ofUO2

2+ and UO2OH+. However, the results fromthe growth inhibition bioassays showed that Uwas more toxic at pH 6.5, corresponding togreater binding of U to the cell surface andgreater uptake of U into the cell. Like Cu, the Utoxicity results may be better explained in termsof competition between H+ and the metal at thecell surface. A number of hypotheses have beenproposed to explain how the H+ in solution offersprotection against metal ions.

The effect of pH on metal toxicity is twofold.The H+ concentration may exert its effect eitherdirectly by affecting metal uptake or indirectly byaffecting the chemical speciation and bioavailabil-ity of the dissolved metal pool (Peterson et al.,1984). Much of the confusion in the literaturesurrounding metal-pH toxicity effects stems froman inability to distinguish between these two fac-tors. Numerous studies have interpreted their re-sults solely in terms of changes in metal speciationin solution, ignoring the possibility of metal–H+

interaction at the cell surface (Starodub et al.,1987). Determination of Cu and U speciation inthis study confirmed that speciation changes wereonly minimal between pH 5.7 and 6.5 (e.g. B10%decrease in Cu2+ from pH 5.7 to 6.5). Also, theuse of a simple inorganic medium minimised theeffects of variables other than pH, such as metalcomplexation, so that changes in toxicityreflected, to a large extent, changes related only toH+ concentration.

Several mechanisms are possible to explain theapparent protective effect of H+. These includecompetition between H+ and the free metal ion atthe cell surface and changes in algal surface po-

tential or membrane permeability. The surfaces ofalgae contain a number of functional groups withhigh affinity for metal ions and carry a net nega-tive charge, mainly due to carboxylic, sulfhydryl,and phosphatic groups (Rai et al., 1981; Crist etal., 1990, 1994). These groups comprise bindingsites that transport metal ions across the cellmembrane and into the cell. It is suggested thatH+ can compete with the free metal ion at the cellsurface for available binding sites. Parent andCampbell (1994) provided direct evidence for H+/Al3+ competition for the green alga, Chlorellapyrenoidosa. In a series of acid titrations, algalcells were shown experimentally to take up moreH+ at pH 4.3 than at 6.5, corresponding toreduced concentrations of cell-bound and intracel-lular Al. These authors, however, concluded thatcompetition at the cell surface could only explainpart of the pH-dependent toxicity.

Changes in algal surface potential have alsobeen implicated as a possible mechanism for in-creased metal toxicity with increasing pH. Crist etal. (1988) demonstrated that as the pH of themedium is increased, an increasing number ofnegatively charged sites are formed on the algalsurface. Because the interaction between metalsand algal surfaces involves electrostatic bonding,this may result in increased metal adsorption tothe cell surface, hence, greater uptake and/or tox-icity. Alternatively, conformational changes inmetal binding sites at the cell surface may occur,thereby altering the cell binding capacity for themetal (Parent and Campbell, 1994).

It has also been suggested that H+ may interactwith the plasma membrane (Macfie et al., 1994),thereby altering its permeability to certain metals,although there is no direct evidence to supportthis hypothesis. Such a mechanism assumes thattransport of the metal across the plasma mem-brane is a prerequisite for toxicity (Parent andCampbell, 1994). While most studies have con-cluded that the mechanism of Cu toxicity to algaeinvolves some form of structural alteration (Wonget al., 1994), or disturbance of the ratio of oxi-dised to reduced glutathione within the cell, lead-ing to reduced growth (Stauber and Florence,1987), it has also been suggested that interferencewith the cell membrane is a fundamental step in

N.M. Franklin et al. / Aquatic Toxicology 48 (2000) 275–289 287

the sequence of Cu2+ toxicity to algae (Gross etal., 1970; Sunda and Huntsman, 1983). Alterna-tively, if we accept that Cu enters the cell byfacilitated diffusion and exerts its toxic effect inthe cytosol or in the cell membrane, the pH effectcould be explained simply by the extent and rateof formation of the Cu-carrier protein(s) com-plex(s) (Florence and Stauber, 1991).

The increased toxicity of Cu and U to Chlorellasp. at pH 6.5 is considered to be a function ofone, or possibly the interaction of all of the abovementioned mechanisms of H+ protection. How-ever, in more complex situations, such as naturalwaters, this relatively simple relationship betweenH+ and metals is likely to be masked by otherrelated factors that also vary with pH, such asmetal complexation with organic matter. Organiccomplexation of metals generally increases withpH, and metal solubility decreases, causing a re-duction in the free ion pool (Peterson et al., 1984).Because of these related factors, the effects ofacidification at the biological surface are some-times found to be less important than are itseffects on metal speciation in solution. Conse-quently, the ability to satisfactorily predict theway in which pH may modify the toxicity ofmetals to biota in natural waters may be limited.Nevertheless, such evidence on pH-dependentmetal toxicity has significant implications for thederivation of appropriate water quality guidelinesfor protecting freshwater biota in tropical waters.

One inherent limitation of this work was thedifficulty in maintaining the pH over the entirebioassay in the poorly buffered test water. Use ofa pH buffer was avoided as some buffers areknown to interfere with the growth of the testspecies and also with metal toxicity (Lage et al.,1996; L. Hales, CSIRO, pers. commun.). An alter-native method of adjusting the pH twice daily wasused, and this proved to be effective in maintain-ing the pH over the first 48 h of the test. How-ever, the effectiveness of this method was largelydiminished over the last 24 h of the test in controltreatments. To overcome this, attempts weremade to shorten the test duration to 48 h. Al-though acceptable control growth rates were ob-tained, the data was more variable over 48 h, andtherefore, a 48-h test was not considered a viable

option. Further work is needed to identify thesuitability of using pH buffers, such as HEPES(pKa=7.5) and MES (pKa=6.1), in metal toxic-ity testing with Chlorella sp.

Acknowledgements

This research was partly funded by the Aus-tralian and New Zealand Environment and Con-servation Council (ANZECC) through theEnvironmental Research Institute of the Supervis-ing Scientist (eriss). The authors would like toMerrin Adams, Michelle Pham, Sherry Yuen, Na-talie Low, Jacquie Lassau, Brett Warden (fromCSIRO Division of Energy Technology) andHenry Wong (Environment Division, ANSTO)for their technical assistance, together with Caro-line Camilleri for use of the Coulter Counter.

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