Speciation, Bioavailability And Toxicity Of Copper In The Fly River System

Discharge of mine tailings into the Fly River since 1984 has resulted in elevated concentrations of particulate and dissolved copper, with documented detrimental impacts on fish populations and other aquatic biota. This chapter reviews recent monitoring of copper and toxicity in the Fly River, assesses recent trends in copper bioavailability, and highlights data gaps in our current understanding of copper effects on biota in the Fly River. While in situ monitoring of fish abundance and species diversity gives an indication of the impact of multiple stressors on fish populations, it does not identify the impact of dissolved copper alone. The most useful data for assessing the effects of dissolved copper come from laboratory-based toxicity tests (bioassays) conducted on aquatic species representative of those found in the Fly River. Several bioassays using copper-sensitive bacteria and microalgae isolated from nearby reference sites, have been used over the past 10 years to monitor copper toxicity in the Fly River system, in conjunction with chemical analyses of dissolved copper, labile copper, and copper-complexing capacity. These bioassays respond to concentrations of biologically available copper in the low micrograms per liter range, and were intended to act as an early warning of copper toxicity to aquatic organisms resident in the Fly River system. Copper speciation, i.e., the chemical forms of copper in natural waters over the pH range 7–8, typical of the Fly River system, is dominated by complexation by natural organic matter. However, in systems receiving elevated inputs of dissolved copper, the complexing capacity of natural organic matter may be exceeded, leading to an increase in the amount of inorganic copper in solution. It is now well established that inorganic copper species, such as the free copper ion or weak or labile complexes that are able to dissociate at the organism–water interface, are more bioavailable than copper in strong or inert complexes or adsorbed to colloidal and particulate matter. The monitoring data indicate that both dissolved copper concentrations and copper-complexing capacities in the Upper and Middle Fly River have remained relatively constant since 1996. However, in the Middle Fly, labile copper concentrations appear to be increasing. The reason for this increase is unknown, but may be due either to inputs of more reactive copper from a change in mining throughput and/or the dredge stockpiles at Bige, or to vegetation dieback on the floodplain and consequent reduced organic matter input or changes in organic matter composition, which in turn influences copper speciation. The frequency of algal growth inhibition has also increased since 1996. Early bioassays with a green alga in 1995 found no toxicity in any of the 10 water samples from sites on the Fly River, despite the fact that they contained between 0.5 and 13 μg/L of dissolved copper, well above the lowest observed effect concentration (LOEC) of 2.5 μg Cu/L for this alga. More bioassays at sites on the Middle Fly River have shown an increase in the frequency of algal growth inhibition, together with bacterial inhibition, corresponding to an increase in labile copper concentrations. The magnitude of algal inhibition, however, has not significantly increased and there is no significant correlation between the amount of algal inhibition and labile copper concentrations in the river. Addition of the metal chelating agent EDTA to water samples consistently removed the observed toxicity in both the algal and bacterial bioassays. This indicates that the toxic effects are likely to be related to the presence of metal contaminants. Of the dissolved metals measured in river waters since 2002, only aluminum and lead exceeded guideline trigger values (ANZECC/ARMCANZ, 2000) on several occasions. Comparison of monitoring data for labile copper to literature data on the sensitivities of freshwater biota to copper shows that algae, invertebrates, insects, and fish are all at risk of chronic effects from copper in the Fly River system. At current labile copper concentrations, chronic effects of copper resulting from long-term exposure to elevated bioavailable copper concentrations may be expected in 50–80% of freshwater species. Any further increases in labile copper in the river may result in chronic toxicity to the majority of freshwater species. In addition, the potential interactive effects of acidification from acid rock drainage (ARD) and metal toxicity are only now being assessed. 10.1 Introduction The Ok Tedi Mining Ltd. (OTML) porphyry gold and copper mine at Mt. Fubilan is an open-pit operation located within the Star Mountains in the Western Province of Papua New Guinea and is situated near the headwaters of the Ok Tedi-Fly River system ( Fig. 10.1 ). Discharge of mine tailings into the Fly River since 1984 has resulted in elevated concentrations of particulate and dissolved copper, with documented detrimental impacts on fish populations and other aquatic biota. Potential impacts of dissolved and particulate copper, released from tailings particles under riverine conditions, are just one of a number of impacts of the OTML mine, including increased turbidity, acidification, channel aggradation, forest dieback, and associated habitat loss. While in situ monitoring, e.g., fish abundance and species diversity, in the Fly River system gives an indication of the impact on populations of these multiple stressors, it does not identify the impact of dissolved copper alone. The most useful data for assessing the effects of dissolved copper come from laboratory-based toxicity tests conducted on aquatic species representative of those found in the Fly River. OTML is currently required to monitor copper bioavailability and aquatic toxicity at four specified locations on the Ok Tedi-Fly River system. The monitoring program is based on the chemical measurement of copper speciation and bioassays using copper-sensitive algae and bacteria, originally intended to provide an early warning of copper toxicity to aquatic organisms resident in the Fly River system. Toxicity tests, together with chemical analyses and biological monitoring, provide several complementary lines of evidence of potential impact of copper in the Fly River system. The accompanying chapter “Biogeochemistry of copper in the Fly River” by Apte (2009) provides a detailed review of historical monitoring data of dissolved copper and copper complexation in the Fly River system. In this chapter, we review speciation data over the last 10 years where labile copper and toxicity in Fly River waters were determined concurrently on the same samples. We also assess recent trends in copper bioavailability in the Fly River system and highlight data gaps in our current understanding of copper effects on biota in the Fly River. 10.2 Overview of Copper Speciation and Bioavailability The potential impacts of copper in aquatic systems depend on the total concentrations of copper, the chemical speciation of copper, interactions of copper at the surface of the organism (e.g., fish gill, algal cell wall), and copper uptake into the organism, with either subsequent adverse effects or intracellular detoxification. Ultimately, adverse effects on cells or individuals may lead to effects at the population or community level. The term ‘bioavailable’ is defined as the fraction of total copper that an organism accumulates, i.e., the fraction of copper that binds to and traverses the cell membrane. The bioavailability of dissolved copper and ultimately its toxicity to aquatic organisms is primarily influenced by the speciation (chemical form) of copper in solution and its binding to receptor sites (e.g., fish gills or algal cell membranes), both of which depend on a range of water quality parameters such as pH, hardness, and dissolved organic matter (DOM) ( Campbell, 1995 ; Markich et al., 2001 ). For copper to be taken up intracellularly where it may exert an adverse effect, it must first bind with ligands present on the organism/cell surface. Lipid bilayer membranes are generally impermeable to charged and polar species, so copper must first be bound to membrane transport proteins that carry copper (and the ligand) into the cell. For copper, the rate of formation of the copper-transport complex is often, but not always, relatively fast compared to copper transport through the membrane, so that a pseudo-equilibrium exists between copper in the external medium and that bound to transport ligands. Thus copper uptake is largely dependent on the free copper ion concentration, which is related to its free-ion activity. In some cases, the uptake of metals may be related to the concentration of kinetically labile inorganic species (free ion plus inorganic complexes) rather than the free metal ion ( Hudson, 1998 ). Such kinetic control of metal uptake may be important when copper transport becomes diffusion limited. Generally, the free copper ion or weak or labile complexes that are able to dissociate at the cell membrane, are more bioavailable than copper in strong or inert complexes or adsorbed to colloidal and particulate matter. There are some exceptions, notably lipid-soluble copper complexes of organic ligands such as xanthates, used as mineral flotation agents ( Florence and Stauber, 1986 ; Stauber and Florence, 1987 ; Phinney and Bruland, 1994 ). These complexes are highly toxic because they can diffuse directly through cell membranes, allowing both copper and the ligand to enter the cell. Although lipid-soluble copper complexes generally represent less than 1% of total copper in natural waters, they may be of greater environmental significance in mine-impacted waters. Over the pH range 7–8, typical of the Fly River system, copper speciation is dominated by complexation by natural organic matter, with copper–organic matter complexes comprising >90% of total dissolved copper concentrations ( Apte and Day, 1993 ). In systems receiving elevated inputs of dissolved copper, the complexing capacity of natural organic matter may be exceeded, leading to an increase in the amount of inorganic copper in solution. In most natural waters, copper toxicity is lower than predicted by the dissolved metal concentration owing to the complexation effects of DOM ( Apte and Day, 1993 ). However, DOM may also exert a direct effect (depending on pH) at the biological surface such as fish gills and algal cells, thereby modifying metal–organism interactions ( Batley et al., 2004 ). Solution pH will also affect copper speciation, with low pH giving rise to larger proportions of the free copper ion. Carbonate concentrations also influence inorganic copper concentrations with copper–carbonate complexes dominating the inorganic copper pool under alkaline solution conditions. The binding of free copper ions to receptor sites on aquatic organisms is influenced by pH, and the concentration of competing cations such as calcium and magnesium (hardness) and sometimes other cations. Unfortunately, it is not possible to generalize on the role of pH and hardness on metal bioavailability. Different organisms have different responses. Although increasing pH can decrease the proportion of free metal ion in solution, there is also a decrease in competing H + , which can actually result in greater copper-cell binding and hence greater copper uptake and toxicity ( Franklin et al., 2000 ; Wilde et al., 2006 ). Traditionally, high hardness has been accepted as having an ameliorating effect on the toxicity of copper and other metals ( Erickson et al., 1996 ); however, recent work in our laboratory has shown that hardness has little protective effect on copper-sensitive freshwater biota ( Markich et al., 2005 ). A range of analytical chemical techniques ( Tessier and Turner, 1995 ) and geochemical modeling approaches ( Allison et al., 1991 ) have been used to measure and predict copper speciation in natural waters. Direct chemical measurement techniques include anodic stripping voltammetry (ASV), ion-selective electrodes, ligand competition methods, ion exchange resins, e.g., Chelex, diffusive gradients in thin films, and size-based separations ( Tessier and Turner, 1995 ; Zhang et al., 1996 ; Batley et al., 2004 ). ASV has been successfully used to detect a labile (inorganic and weakly bound organic) copper fraction and to determine copper complexation capacity. While the labile copper measured at natural pH is operationally defined by the measurement technique, it is generally thought to be related to bioavailable copper. There are exceptions, particularly when waters contain high DOM ( Stauber et al., 2000 ). ASV-labile copper in the Fly River samples was measured using a Metrohom 746VA trace Analyser and a hanging mercury drop electrode, according to the method outlined in Apte et al. (2005) . PTFE (Teflon) voltammetric cells were pre-equilibrated with the river water for at least 30 min, before being discarded. Buffered test solutions (20 mL) were weighed into the PTFE cells and sodium nitrate added (40 μL of 5 M). ASV-labile copper was determined after purging the samples for 5 min with nitrogen. Geochemical speciation models have been used for many years to calculate copper speciation at a range of water quality characteristics ( Turner et al., 1981 ; Soli and Byrne, 1989 ). Most assume chemical thermodynamic equilibrium and until recently, their use has been limited by the lack of known binding constants for natural DOM and inability to account for metal adsorption. Franklin et al. (2000) showed that although copper speciation changes in a synthetic soft water, as calculated using geochemical speciation modeling, were negligible over the pH range 5.7–6.5, the toxicity of copper to a freshwater green alga varied by more than a factor of 20, indicating that speciation modeling may have limited ability in predicting copper bioavailability to aquatic organisms. Moreover, speciation models based on chemical thermodynamic equilibrium cannot predict metal uptake that is under kinetic control. 10.3 Bioassays While chemical measurement techniques and geochemical speciation modeling may detect or predict the different forms of copper in aquatic systems and hence give some information about likely toxicity, they do not provide direct data on adverse biological effects. Bioassays or toxicity tests are generic tests that use living organisms as indicators of contaminant bioavailability in aquatic systems. Acute bioassays (short-term tests) typically measure organism survival over 96 h or a sublethal effect such as bioluminescence or enzyme inhibition. Chronic tests, such as inhibition of growth of microalgae, determine toxicity over several generations of cells. Such tests may be of long duration (weeks) or short-term in the case of single-celled algae (that divide approximately once per day). Copper may cause a decrease in final cell numbers, a decrease in exponential growth rate, or an increase in the time to commencement of growth (lag phase). Because different organisms have different sensitivities to copper, batteries of toxicity tests, using sensitive species from different trophic levels, are usually used. Several bioassays using copper-sensitive bacteria and microalgae isolated from reference sites in Papua New Guinea, have been used over the past 10 years to monitor copper toxicity in the Fly River system, in conjunction with chemical analyses of dissolved copper, labile copper, and copper-complexing capacity. These bioassays respond to concentrations of biologically available copper in the low micrograms per liter range. These bioassays, developed specifically for application in the Fly River, are described in more detail below. 10.3.1 Bacterial Bioassays Bacteria have several attributes that make them useful as test organisms for the screening of metals in natural waters. They have relatively short life cycles, respond quickly to environmental change, and have enzymatic processes common to those of higher organisms. Bacterial processes are also of vital importance in the aquatic environment, mediating the degradation of organic compounds, biogeochemical cycling of nutrients, and transformations of trace metals ( Lee et al., 1990 ). The response of bacteria to contaminants such as copper ranges from relative tolerance to extreme sensitivity, with some species exhibiting sensitivity at near-ambient metal concentrations. Sensitive bacterial bioassays were developed with an isolate from the Fly River, upstream from its confluence with the Ok Tedi ( Davies et al., 1998 ), and this was used in early monitoring of copper toxicity in the Fly River system in the mid-1990s. However, problems with maintaining consistent sensitivity of the bacterium in culture to copper necessitated the isolation of a temperate species ( Erwinia persicinus ) from Sydney for subsequent toxicity monitoring. These bioassays, developed in our laboratory, are described below. 10.3.1.1 Growth inhibition bioassay The bacterial growth inhibition test utilizes a copper-sensitive freshwater bacterium, E. persicina , isolated from the Woronora River, New South Wales ( Rogers et al., 2005 ). The bioassay is highly reproducible and exhibits a sensitivity to copper at <2 μg/L. The inhibition of cell division (growth) is measured over 48 h in various dilutions of freshwater samples supplemented with nutrient media, and compared to a pristine control water and a solution of known copper concentration. Growth is determined by optical density measurements (420 nm) 48 h after inoculation. The concentration of the test sample causing a 50% inhibition in bacterial growth and that at which no effect is observed are determined using standard statistical methods. Criteria for test acceptability include a final optical density at 420 nm of 0.20±0.05 for the control sample, a statistically significant reduction in growth in the presence of 2 μg/L copper, and less than 20% coefficient of variation in the controls. This type of bacterial growth test (with another species) has been used to examine the toxicity of copper in a number of freshwaters ( Davies et al., 1998 ). The sensitivity of the bacterial isolate to other metals in unknown. 10.3.1.2 Radiochemical bioassay The uptake and assimilation of radiolabeled metabolites such as glucose by bacteria can also be used to quantify microbial responses to metal contamination ( Gillespie and Vaccaro, 1978 .). A radiochemical freshwater bioassay has been developed using the same metal-sensitive bacterial isolate used in the growth inhibition test ( Rogers et al., 2005 ). A 50% reduction in the assimilation of radiolabeled glucose is observed at <1 μg/L copper, which corresponds well with the growth bioassay response. Due to the sensitivity of radiochemical techniques, it is possible to use environmentally relevant bacterial concentrations and short incubation periods thus minimizing the potential for changes in metal speciation during the bioassay. Furthermore, metal-impacted waters may be tested without the need to add growth-stimulating nutrients. A bacterial culture is grown in a defined medium for 24 h. The cells are harvested by aseptic centrifugation, washed twice, and resuspended in sterile synthetic water. Test solutions are inoculated with the washed starter culture to achieve a final cell density of 10 5 cells/mL. The inoculated test solutions are equilibrated for 2 h followed by addition of an aliquot of d -glucose-UL- 14 C solution. A further 2 h is allowed for cellular glucose assimilation and then formalin is added to terminate respiratory activity. The cells are harvested by filtration and the β-activity associated with each sample is measured using a liquid scintillation counter. The assimilation of 14 C-glucose by copper-exposed cells is expressed as percentage assimilation compared to the control. Criteria for test acceptability include an initial optical density (420 nm) between 0.1 and 0.2 (corresponding to a cell density of 1.4±0.2×10 7 cells/mL), <20% coefficient of variation in controls and a statistically significant reduction in glucose assimilation in the presence of 1 μg/L copper compared to the control. 10.3.2 Algal Bioassays Microalgae are particularly important in tropical aquatic ecosystems, being responsible for most of the primary production at the base of the aquatic food chain. Recent stable-isotope studies have clearly shown the importance of microalgae in the food webs of the Fly River system Storey (2005) . A bioassay developed in our laboratory measures the decrease in cell division rate (growth rate) of the unicellular alga Chlorella sp. in a 3-day exposure to toxicant. This species was originally isolated from Lake Aesake, Papua New Guinea. It is cultured axenically on a 12:12 h light/dark cycle (cool white fluorescent light, 100 μmol photons m −2 s −1 ) at 27°C. Cells in log-phase growth are used in the algal bioassays after washing and centrifuging three times to remove culture medium. The bioassay follows the OECD Guideline 201 ( OECD, 1984 ) and the protocol of Stauber et al. (1994) . Controls in synthetic soft water, together with Fly River water samples, each in quadruplicate, are prepared. Fifty milliliters of each is dispensed into 200 mL silanized (Coatasil, BDH) Erlenmeyer flasks. To each flask, 0.5 mL of 26 mM sodium nitrate and 0.05 mL of 1.3 mM potassium dihydrogen phosphate is added as nutrients. Each flask is inoculated with 10 3 or 10 4 cells/mL of a prewashed algal suspension and incubated at 27°C on a 12:12 h light/dark cycle at 100 μmol photons m −2 s −1 for 72 h. The pH in each flask is measured on Day 0 and Day 3 of the bioassay. Cell densities in each flask are determined daily for 3 days by counting cells in either a Coulter Multisizer II particle analyser with 70 μm aperture or a flow cytometer. A regression line is fitted to a plot of log 10 cell density versus time ( h ) for each flask and the cell division rate (growth rate) per hour ( μ ) determined from the slope. Cell division rates per day (3.32× μ ×24) are calculated for each water sample and expressed as a percentage of the control growth rate. The test is considered acceptable if the algal cell division rate in the soft water controls is 1.4±0.4 doublings/day and the variability in the controls is <20%. The reference toxicant copper (tested at at least three concentrations) is included in the bioassay to ensure that the algae are responding to a known toxicant in a reproducible way. The 72-h IC 50 (i.e., the inhibitory concentration to cause a 50% decrease in algal cell division rate compared to controls) is calculated and the bioassay is acceptable if the copper IC 50 is within quality control chart limits. 10.3.3 Limitations of the Algal and Bacterial Bioassays While bioassays have a number of advantages over other monitoring techniques such as the ability to detect bioavailability of mixtures of contaminants, high sensitivity, ecological relevance, and reproducibility, they also suffer from some limitations ( Stauber and Davies, 2000 ). In particular, bioassays can underestimate copper toxicity if copper in the test solutions is depleted over the duration of the test, whether by adsorption to cells or production of cell exudates, which bind copper and reduce its bioavailability. Franklin et al. (2002) showed that copper toxicity to Chlorella sp. 12 decreased with increasing initial cell density from 10 2 to 10 5 cells/mL. With the recent application of flow cytometry to algal ( Franklin et al., 2000 ) and bacterial ( Boswell et al., 1998 ) toxicity testing, there is now the capability to conduct bioassays at much lower cell densities (down to 10 2 cells/mL for algae) that are more typical of environmental concentrations. Such low cell densities have been used for toxicity testing of Fly River waters since 2004 to improve the environmental relevance of the bioassays. Different organisms respond differently to contaminants such as copper, hence the need for a range of toxicity test species. Recent work in our laboratory has shown that the effects of copper on Chlorella sp. were ameliorated in natural water samples, i.e., less toxic than predicted based on the free copper ion concentration, whereas the bioavailable copper fraction for bacteria was much greater than the free metal alone ( Apte et al., 2005 ). A possible explanation for the observed amelioration of algal toxicity in the presence of natural DOM is the adsorption of DOM to the surface of the alga and the consequent blocking of metal receptor sites. In contrast, the bacteria could have responded to some forms of organically complexed copper, which were bioavailable, hence copper toxicity may be greater than that predicted by the free metal ion. Thus the two different bioassays appear to be measuring different fractions of bioavailable copper. 10.4 Historical Toxicity Monitoring Data in the Fly River System Few data have been collected on the direct toxicity of copper on aquatic organisms indigenous to the Fly River system. The two most studied groups of aquatic organisms in the Fly/Strickland system are algae and large fish, with only limited data available on invertebrate species, despite the fact that they make up a significant proportion of the aquatic fauna. There are no data available, at this time, on the toxicity of copper or Fly River water to planktivorous fish. 10.4.1 Invertebrate and Fish Bioassays Laboratory-based ecotoxicological tests of OTML mining wastes have been reviewed by Smith (1997) . Toxicity testing using fish, prawns, cladocerans and mayflies showed that these species had sensitivities to dissolved copper comparable to published sensitivities for Australian species. Smith et al. (1990) reported the toxic effects of particulate copper to the freshwater prawns Macrobrachium rosenbergii and Macrobrachium handschini , and the catfish Neosilurus ater . The test medium was synthetic Fly River water reconstituted from Woronora River (NSW) water with the addition of sodium bicarbonate to control pH and alkalinity. If the dissolved copper concentrations were kept below those likely to occur in the Fly River, little impact from particulate copper was observed at concentrations ranging from 14,300 to 15,100 μg Cu/g, much higher concentrations than those expected to occur at sites downstream from the mine. Recent studies of acute copper toxicity to juveniles (mean length 306±21 mm) of a native barramundi species in the Fly River, reported 96-h LC 50 values between 410 μg/L (initial measured concentration) and 270 μg/L (final measured concentration) dissolved copper at pH 7.0–7.6 and 20.5 mg/L CaCO 3 ( Australian Water Technologies, 2002 ). The dilution water used for the test medium was carbon-filtered Melbourne mains water, which would not have been representative of the natural water in the Fly River system. 10.4.2 Microalgal Bioassays In terms of copper ecotoxicology, microalgae represent the most studied group of aquatic organisms in the Fly River. It is clear that indigenous algal species are sensitive to relatively low concentrations of bioavailable copper. Early bioassays to investigate the potential toxicity of copper in the waters of the Fly River used a temperate green algal species Chlorella protothecoides obtained from the CSIRO Division of Fisheries Culture Collection, Hobart ( Stauber and Critelli, 1993 ). These tests compared the growth of C. protothecoides in pristine laboratory water controls, matched to the same hardness, pH, nitrate and phosphate concentrations as Fly River water, to the growth of this alga in Fly River water at pH 7.9. Eight river water samples from Ok Tedi, Ok Mani, and the Fly River, together with two off-river water bodies (Bosset Lagoon and Lake Daviumbu) were tested ( Fig. 10.1 ). No toxicity was observed for any of the 10 water samples from sites on the Fly River, despite the fact that they contained between 0.5 and 13 μg/L of dissolved copper, well above the lowest observed effect concentration (LOEC) of 2.5 μg Cu/L for this alga. Copper speciation measurements were not made on these early samples. In 1995, Stauber carried out a brief taxonomic survey of algal species present in the Fly/Strickland River system, with the aim of isolating suitable copper-sensitive species for use in the development of an algal growth inhibition bioassay with local tropical species. Tropical algae from eight sites in the Fly and Strickland River systems were collected in January 1995. Two sites, Lake Daviumbu and Lake Pangua in the Fly River floodplain, were potentially impacted by copper, while six reference sites in the Fly River oxbows near Kiunga or near the Strickland River (Lake Aesake and Oxbow Levamme) were not copper impacted. Good species diversity was found in all lakes at the time of this survey. Over 60 algae were identified to genus level, with the dominant taxa at all sites being green algae (Chlorophyceae), diatoms, and blue-green algae (Cyanophyceae) ( Stauber and Apte, 1996 ). There was little difference in the genus composition or diversity between the impacted lakes in the Fly River system and the reference lakes in the Strickland River system. Thirty-five genera of green algae were identified, of which nine were only present in nonimpacted waters and two were present only in copper-impacted waters. The filamentous blue-green algae were distributed mainly in Lake Daviumbu, Lake Pangua, and Lake Aesake and were typical of the flora of shallow tropical water bodies. Of the 14 genera of diatoms identified, only 2 ( Cocconeis and Cyclotella ) were present exclusively in the Strickland River reference sites. The distribution of chrysophytes between samples from the Fly River and Strickland River systems was very different, although few genera were found. Four of the five genera identified were only found in nonimpacted sites and one genus ( Bicosoeca ) was found only in impacted lakes. In general however, the flora of both the Fly River and Strickland River waters were similar, and broadly similar to algae of the Sepik River floodplain in Papua New Guinea (except for the absence of desmids) (W. Vyverman, personal communication). The diversity of euglenoids found (five genera) was typical of mixed water lakes that receive sediment-loaded water during the wet season and flood plain-derived water during the dry season. Dominant algal genera in phytoplankton net samples during this 1995 survey are shown in Table 10.1 . Algal abundance was also determined by cell counts in bottle-collected samples. All samples from both impacted and nonimpacted sites contained large numbers and diversity of blue-green algae, diatoms, dinoflagellates, green algae, and a cryptomonad (1–640 cells/mL). Stauber and Apte (1996) isolated 21 strains of microalgae, comprising 9 species, from reference sites in the Fly/Strickland River system. Most of these were green algae from Lake Aesake, together with two diatoms ( Nitzschia palea and Achnanthes sp). Unialgal (and some bacteria-free) cultures were established, and taxonomically identified. All isolated species were ubiquitous, widely distributed and identified in the original field samples. Nine isolates of three common species ( Chlorella , Chlamydomonas , and Monoraphidium arcuatum ) were also tested for copper sensitivity in screening bioassays. All species were sensitive to copper, with 72-h IC 50 values ranging from 7 to 17 μg Cu/L, LOEC values of <10 μg Cu/L and NOEC values of <5 μg Cu/L. A copper-sensitive green alga, Chlorella sp. 12, isolated from Lake Aesake on the Strickland River, was selected and subsequently used in the development of the site-specific growth rate inhibition bioassay, described earlier. Chlorella sp. 12 is particularly sensitive to copper, with copper toxicity depending on the initial cell density used in the bioassay ( Franklin et al., 2002 ) and water quality parameters. For the standard test that used 2–4×10 4 cells/mL as the initial inoculum in Fly River monitoring up until 2004, the 72-h IC 50 (mean±2SD) was 8.5±4.6 μg Cu/L, with a no observed effect concentration (NOEC) at 5 μg Cu/L. More recent tests using lower cell densities (2–4×10 3 cells/mL) to improve environmental realism and sensitivity, have given slightly lower IC 50 values of 5.5±4.5 μg Cu/L, with a LOEC at 4 μg Cu/L and NOEC at 3 μg Cu/L. This is of comparable sensitivity to copper as the temperate C. protothecoides used in the early bioassays of Fly River waters, suggesting that any changes in copper toxicity over time were unlikely to be due to changes in bioassay procedures. Recent work has shown that copper toxicity to Chlorella sp. 12 is highly dependent on water quality characteristics ( Wilde et al., 2006 ). Copper toxicity in synthetic soft water decreased about 20-fold as the pH decreased from 8.0 to 5.5, and decreased about 80-fold as the DOC increased from 0 to 20 mg/L. Interestingly, hardness had no effect on copper toxicity to Chlorella sp. 12 in the synthetic water bioassays. A comparative study ( Stauber and Apte, 1996 ) used C. protothecoides, Chlorella sp. 12, a bacterial bioassay ( Apte et al., 1995 ; Davies et al., 1998 ), and ASV to measure the complexation capacity of water samples from seven sites on the Fly River, which had been collected over the preceding 12-month period. To measure complexing capacity from bioassays, water samples were spiked with additional copper, and assessed for toxicity in the usual way. The IC 15 (i.e., the concentration of copper to cause a 15% reduction in growth rate) was taken as a measure of the copper-complexing capacity of the sample. All of the samples had copper complexation capacities (4–33 μg/L) in excess of their respective dissolved copper concentrations (1.8–17 μg/L), and again no toxicity was observed at any of the sites ( Table 10.2 ). In fact, significant growth enhancement of Chlorella sp. 12 was observed in the four samples tested with this species compared to controls. 10.4.3 Monitoring of Copper Speciation and Toxicity using Anodic Stripping Voltammetry and Bacterial and Algal Bioassays Total dissolved copper concentrations and copper speciation data, in conjunction with ecotoxicological bioassay data, have been periodically collected by OTML and CSIRO since 1993. The aim was to gain temporal as well as spatial information on copper speciation and bioavailability, and most importantly, to link the chemical form of dissolved copper with biological measurements of toxicity. Early surveys in 1993 and 1995, using C. protothecoides and Chlorella sp. 12, together with a bacterium isolated from the Upper Fly River (Isolate 37), found no toxicity at any sites in the Fly River, despite the fact that dissolved copper concentrations (0.5–17 μg/L) exceeded the test species LOEC values for copper. The absence of toxicity was attributed to DOM complexation lowering copper bioavailability. There were considerable logistical difficulties in the collection and transport of Fly River waters for testing in our Sydney laboratories. To avoid contamination, water samples had to be filtered on arrival in Sydney, rather than on-site. Maintaining the temperature of the samples at 4°C in transport was also difficult, yet holding temperature was found to be a critical factor in determining the amount of labile copper in the samples. Labile copper decreased markedly within a few weeks of storage of the water samples at room temperature, so it is possible that, on occasions, particularly in the early survey work, labile copper and hence toxicity was underestimated. 10.4.3.1 Speciation/toxicity survey 1996–1997 A more detailed study of copper speciation and copper toxicity was carried out over the period September 1996–June 1997 using both electrochemical speciation techniques and copper-sensitive algal and bacterial growth bioassays ( Apte et al., 1997 ). A total of 48 river water samples in 7 surveys were analyzed for total dissolved copper, ASV-labile copper, copper complexation capacity, and algal and bacterial growth inhibition. Total dissolved copper concentrations ranged from 5.4 to 42.6 μg/L at sites downstream from the mine (41 samples) and from 1.0 to 3.8 μg/L at the riverine reference site Kiunga. The highest dissolved copper concentration was measured at Ningerum, the site closest to the mine. ASV-labile copper (0.5–8.3 μg/L) was detected in 26 of the 41 samples, but was not detected at the reference site. In eight of the samples, labile copper was detected even though the copper complexation capacity had not been exceeded. It was suggested that this was due to the presence of weakly bound copper–organic complexes which dissociated during the ASV analysis. Algal and/or bacterial growth inhibition was observed at four out of the seven sites sampled (at Ningerum six times and, D’Albertis Junction, Nukumba, and the Fly at Bosset each on one occasion). Samples from Ningerum were consistently toxic on six out of eight occasions. For both the algal and bacterial bioassays, dissolved copper in the river water samples was less toxic than for an inorganic copper calibration bioassay carried out in laboratory water, illustrating the role of the natural organic matter present in the natural waters in ameliorating copper toxicity. There was no general relationship between algal growth inhibition and labile copper, but inhibition was only observed when labile copper was present. These data could be split into two populations ( Fig. 10.2 ): 1. No toxicity observed in the presence of labile copper. For these samples, labile copper concentrations were generally below the LOEC for inorganic copper in synthetic water (about 5 μg Cu/L) for Chlorella sp. 12, i.e., there was insufficient labile copper to cause any growth inhibition. 2. More toxic than predicted by the labile copper concentration. In these samples, the concentrations of copper to cause growth inhibition were typically 5 μg Cu/L lower than the concentrations in the inorganic copper calibration bioassays. The most likely explanation for this was the presence of easily dissociable copper–organic complexes, or the presence of lipid-soluble copper complexes which have a greater toxicity than ionic copper ( Florence and Stauber, 1986 ). Alternatively, other toxic metals may be present in the sample, i.e., the growth inhibition observed was not solely due to copper. In the final survey conducted in June 1997, consistent trends in copper speciation and algal and bacterial growth inhibition were observed. Labile copper was detected at three sites closest to the mine input (Ningerum, D’Albertis Junction, and Nukumba) and both bioassays showed growth inhibition at these sites. The linear relationship between algal growth inhibition and labile copper concentration ( m =12.8±2.4, r 2 =0.967) was similar to that for ionic copper in the calibration bioassay ( m =16.3±0.2, r 2 =0.999). Less labile copper was needed to elicit a similar degree of growth inhibition, the labile copper curve being displaced from the ionic copper calibration curve by about 5 μg/L. Again, this suggested that a fraction of the copper–organic complexes were bioavailable to the algae. Finally, copper uptake experiments were performed using three of the samples. Cellular copper concentrations increased with increasing amounts of labile copper in the samples, and both the algae and bacteria showed the greatest amount of copper uptake from the Ningerum sample. 10.4.3.2 Speciation/toxicity surveys 2004-onwards Since early 2004, monitoring surveys have analyzed a total of 53 water samples for labile copper and toxicity to bacteria ( E. persicina ) and microalgae ( Chlorella sp. 12). To avoid the confounding factor of algal growth or bacterial respiration stimulation due to nutrients in the water samples, matrix-matched controls were prepared by the addition of the chelating agent EDTA to one subsample of each of the test waters ( Apte et al., 1997 ). EDTA complexes metals such as copper, rendering them nontoxic. Microalgal growth or bacterial respiration in the EDTA-amended samples was then compared to growth/respiration in the non-amended samples. In this way, each water sample served as its own control. Of the 53 water samples tested, toxicity has been detected by either or both the algal or bacterial bioassays in 38 samples (72%). The sites and frequency at which toxicity was observed are shown in Table 10.3 . Nine different sites were tested and toxicity was observed at seven sites (shown in Table 10.3 ). In general, toxicity was observed consistently at sites on the Upper Fly River and only intermittently at sites on the lower river or in off-river water bodies. No toxicity was observed at the control site Kiunga (two samples) or at Oxbow four (two samples) (not shown). This suggests that the number of potentially toxic samples has increased compared to the 1996–1997 surveys. Good agreement was obtained between the algal and bacterial bioassays during this period, with differences occurring only for 7 of the 53 samples. Although there was again no relationship between dissolved or labile copper and toxicity, labile copper was always present when toxicity was observed. For the microalga, no clear relationship was discernable between dissolved copper and algal growth inhibition ( Fig. 10.3a ). This plot also includes the algal bioassay response to inorganic copper in laboratory synthetic freshwater, represented by the inorganic copper calibration curve. As in the previous surveys, the concentration of copper in the test waters required to elicit an inhibitory response was much higher than for inorganic copper (i.e., to the right of the inorganic copper calibration curve). Eight test waters exhibited toxicity in the region of the calibration curve, suggesting that there was little ameliorative effect of DOM in these eight waters. The relationship between algal growth inhibition and labile copper measured by ASV is shown in Fig. 10.3b . Four regions are discernable. First, concentrations of labile copper in the samples that were not inhibitory to algal growth were generally at or below 6 μg/L, close to the threshold lowest observable effect concentration for ionic copper for this bioassay. Therefore insufficient labile copper was present to cause an inhibitory effect. Second, some of the waters tested exhibited toxicity in the region of the inorganic copper calibration curve suggesting that the measured labile copper was responsible for the observed toxicity. The rest, and majority, of the data fell either to the right of the calibration curve, indicating that the measured labile copper was less toxic than predicted, or to the left of the curve indicating greater than predicted toxicity. The latter may be due to the presence of additional metals or other toxicants in the water samples. Addition of the metal chelating agent EDTA to water samples consistently removed the observed toxicity, indicating that the toxic effects are most likely related to the presence of metal contaminants. These may be independently toxic, or have an additive, antagonistic or synergistic effect on copper toxicity to Chlorella sp. 12. Monitoring of other metals in the water samples, including aluminum, cadmium, chromium, iron, manganese, nickel, lead, and zinc, has shown that lead exceeded its guideline trigger value ( ANZECC/ARMCANZ, 2000 ) of 3.4 μg/L at all sites and aluminum exceeded its guideline trigger value of 55 μg/L at Ningerum and Nukumba (on two occasions). There was no relationship between the concentrations of lead (11–12 μg/L at all sites on one occasion it was measured) and the observed toxicity, suggesting that lead was not responsible for the additional toxicity. In contrast, samples containing slightly elevated aluminum (58 and 56 μg/L) were more toxic than predicted from their labile copper concentrations, suggesting that aluminum may contribute to toxicity. However, the toxicity of aluminum alone to Chlorella sp. has not been determined, so it is difficult to assess whether this may have contributed to the algal growth inhibition on several occasions. For the bacteria, increasing dissolved copper concentrations generally increased inhibition in the bacterial bioassay ( Fig. 10.4a ). This plot also includes the bacterial bioassay response to inorganic copper in laboratory synthetic freshwater, represented by the inorganic copper calibration curve ( Rogers et al., 2005 ). In many of the test waters, no inhibitory response was observed, despite the presence of dissolved copper concentrations up to 10 μg/L, illustrating the role of dissolved natural organic matter in ameliorating bacterial copper toxicity. However, the concentration of copper in the test waters required to elicit an inhibitory response was much higher than for ionic copper (i.e., to the right of the inorganic calibration curve). The relationship between bacterial inhibition and labile copper measured by ASV is shown in Fig. 10.4b . Increasing concentrations of labile copper were correlated with inhibition of the bacterial bioassay ( R =−0.703). However, despite the presence of ASV-labile copper concentrations between 1.3 and 3.5 μg/L in six of the test waters, no bacterial toxicity was observed. 10.5 Spatial and Temporal Trends in Copper Toxicity in the Ok Tedi/Fly River System It is difficult to determine temporal trends in labile copper and toxicity, as previous studies in which dissolved copper concentrations were intensively monitored at Nukumba and Bige ( Shao et al., 2002 ), indicated large inherent variability in dissolved copper concentrations. With such large variability in dissolved copper at hourly, daily, weekly, and monthly timescales, it is not possible to draw firm conclusions about long-term trends in copper bioavailability. It is likely that the greatest copper toxicity is observed after events such as drying and wetting of the floodplain (e.g., the first rainfall after a prolonged dry period). Event-based sampling is therefore required to better interpret trends. In addition, acidification of the Fly River system, first detected in 2000, confounds interpretation of toxicity changes over time, as pH and copper can interact to have additive or antagonistic effects on biota. Nevertheless, general comments on labile copper and toxicity data at four main sites in the Fly River system monitored since 1996 are given below. 10.5.1 Ok Tedi Dissolved copper concentrations at Ningerum ranged from 6 to 23 μg/L but appear to have decreased in the later survey period ( Fig. 10.5a ). Concentrations from April 2002 to July 2004 (6–7 μg/L) were significantly ( p =0.02) lower than from September 1996 to February 2002 (11–23 μg/L). Labile copper concentrations at Ningerum were generally low (<4 μg/L) during the 1996–1997 survey period, although algal toxicity was observed on three occasions ( Fig. 10.5b ). A small increase in labile copper was observed in the period February 2002 to April 2003, with measured concentrations being consistently above 4 μg/L and rising to a maximum of 9.2 μg/L. This was accompanied by consistent algal growth inhibition. Labile copper concentrations have subsequently decreased to between 3 and 5 μg/L but this has not been accompanied by a concurrent decrease in algal growth inhibition at this site. There is a weak correlation ( R =0.657) between the labile copper concentration and the extent of algal growth inhibition at Ningerum. 10.5.2 Middle Fly River At Nukumba, dissolved copper concentrations were similar to those observed at Ningerum and ranged from 9 to 23 μg/L ( Fig. 10.6a ). At this site, labile copper concentrations were <4 μg/L during the early survey period, November 1996–June 1997. Algal inhibition was observed on only one occasion but was accompanied by the lowest measured labile copper concentration (0.6 μg/L) ( Fig. 10.6b ). Labile copper concentrations increased during the period February 2002–February 2003 being consistently above 5 μg/L and rising to a maximum of 9.3 μg/L, but algal growth inhibition was not always observed. Labile copper concentrations appeared to decrease at Nukumba from April to December 2003, but algal growth inhibition was still consistently observed. Overall there was a very poor correlation ( R =0.394) between algal growth inhibition and labile copper concentrations at Nukumba. At Obo, dissolved copper concentrations (11–28 μg/L) and copper complexation capacities (9–38 μg/L) remained high and were comparable to those observed at sites on the upper river. Labile copper concentrations were low (<0.5–8 μg/L) ( Fig. 10.7a ). However, there has been a general, and statistically significant ( p <0.001), increase in labile copper concentrations between the 1996/1997 surveys and the 2002/2004 surveys. No algal inhibition was observed at Obo during the 1996/1997 surveys when labile copper concentrations ranged from <0.5 to 2.5 μg/L. After February 2002, labile copper concentrations were consistently >3 μg/L, reaching a maximum value of 8.2 μg/L. Algal inhibition was also observed at Obo from February 2002 to July 2004 ( Fig. 10.6b ) but there was only a poor correlation ( R =−0.317) between labile copper concentrations and the magnitude of algal inhibition during this latter survey period. Lower dissolved copper concentrations (5–14 μg/L) were observed at Ogwa compared to sites further up the Fly River ( Fig. 10.8a ). Labile copper concentrations however, showed a similar trend to those observed at Obo. During the early survey period, September 1996 to June 1997, labile copper concentrations were <1 μg/L and no algal inhibition was observed. Increased labile copper concentrations ranging from 2 to 4 μg/L were observed from August 2003 and consistent algal inhibition was observed ( Fig. 10.8b ). Further work is required to determine if the recent increase in algal inhibition at Ogwa is significant, and if there is any correlation between increasing labile copper concentrations and the extent of algal growth inhibition observed at this site. The poor correlations between labile copper concentration and algal growth inhibition, and the lack of an obvious relationship between dissolved copper concentrations, copper complexation capacities, and labile copper concentrations at these sites, strongly suggests that factors such as the concentrations of other dissolved metals, e.g., aluminum, or suspended particulate matter in the river may be influencing the observed algal toxicity. 10.6 Risks to Aquatic Biota in the Fly River System The speciation monitoring data indicate that both dissolved copper concentrations and copper-complexing capacity in the Upper and Middle Fly River have remained relatively constant since 1996. Although there were some logistical difficulties in the early speciation surveys, which may have resulted in an underestimation of labile copper and copper toxicity, there is a general trend of increasing labile copper concentrations in the Middle Fly. The reason for this increase is unknown, but may be due either to inputs of more reactive copper from a change in mining throughput and/or the dredge stockpiles at Bige, or to vegetation dieback on the floodplain and consequent reduced organic matter input or changed organic matter composition, which in turn influences copper speciation. The frequency of algal growth inhibition has also increased since 1996. At sites on the Middle Fly River, the increase in the frequency of algal growth inhibition corresponds to an increase in labile copper concentrations. The magnitude of algal inhibition however, has not significantly increased and there is no significant correlation between the amount of algal inhibition and labile copper concentrations in the river. Addition of EDTA to water samples consistently removed the observed toxicity in both the algal and bacterial bioassays. This indicates that the toxic effects are likely to be related to the presence of metal contaminants. Of the dissolved metals measured in river waters since 2002, only aluminum and lead exceeded guideline trigger values ( ANZECC/ARMCANZ, 2000 ) on one or more occasions. In an attempt to understand the impacts of copper toxicity on the Fly River aquatic ecosystem, a summary of the mean and maximum labile copper concentrations at four sites in the Ok Tedi and Fly Rivers is given in Table 10.4 , together with the cumulative frequency distribution plot of labile copper at all monitoring sites in the Ok Tedi and Fly Rivers over the recent monitoring campaign ( Fig. 10.9a ). A comparison of Fig. 10.9a monitoring data for labile copper with known sensitivities of freshwater biota to dissolved copper in the literature [ Fig. 10.9b , summarized in Rogers et al. (2005) and Parametrix (1999) ] shows that freshwater biota are at risk of chronic effects from copper in the Fly River system. At the labile copper concentrations measured during the monitoring program, acute copper toxicity is not an issue. At worst, acute toxicity may be observed in fewer than 10% of invertebrate species, with no acute effects on fish. However, at current labile copper concentrations, chronic effects of copper resulting from long-term exposure to elevated bioavailable copper concentrations may be expected in 50–80% of freshwater species. Any further increases in labile copper in the river may result in chronic toxicity to the majority of freshwater species ( Rogers et al., 2005 ). Food web stable-isotope studies ( Storey, 2005 ; Storey and Yarrao, 2009 ) clearly indicate the importance of microalgae, such as periphyton, in the food webs of the Fly River system. Studies by Bunn et al. (1999) indicate that about 40% of the riverine and 70% of the floodplain fish biomass is supported by carbon derived from algae. Recent work by Storey (2005) , suggests that periphyton abundance is reduced in the river sections downstream of D’Albertis Junction, and the contribution of algal carbon to the aquatic food web in these sections is also reduced compared with areas upstream of the mine. The factors affecting periphyton abundance in the Fly River remain to be established. At this stage, it is not possible to deconvolute the effects of copper, turbidity (with respect to light production or scouring) and loss of habitat. Nevertheless, the concentrations of labile copper observed in the river system are in the range that may cause growth reduction in algae. Loss of algal carbon food sources and consequent effects on higher-level consumers, irrespective of the cause, is an important potential impact. 10.7 Looking Ahead While the labile copper and toxicity monitoring data over the past 10 years have provided valuable information on spatial and temporal trends in copper in the Fly River system, further event-based sampling will enable the assessment of variability in the system and better analysis of long-term trends. Changing mine practices and increased forest dieback currently confound our understanding of the causes of the observed toxicity and we have a very limited understanding of the effects of colloidal copper (included in the estimates of “dissolved” copper) or other metals on biota in the Fly River. The potential interactive effects of acidification from acid rock drainage (ARD) and metal toxicity have not been assessed, but have the potential for serious consequences if pH reductions are significant. While ecotoxicological studies under controlled conditions with algae isolated from the Fly River have shown effects on algal growth in Fly River water, extrapolating this response to the response of complex aquatic communities such as those found in the Fly River is difficult. In particular, effects such as ecosystem recovery, depletion of food sources, prey switching, and algal succession are not considered in laboratory studies. In a recent review by Chariton and Apte (2005) , they concluded that, while copper concentrations below 2 μg/L have no effect on composition, diversity, and abundance of biota, at slightly higher concentrations (4 μg Cu/L) changes in algal assemblages, including the dominance of taxa, have been demonstrated. Concentrations of copper above 10 μg/L can lead to a reduction in algal biomass and a shift to copper-tolerant species. Reduced pH (<6.2) can independently lead to an imbalance in the ratio of algal production to respiration, resulting in the dominance of acid-tolerant diatoms. Combined low pH (<5) and elevated aluminum concentrations also have the potential to impact the Fly River system in future. The presence of localized ARD on levee banks of the Middle Fly, resulting in elevated dissolved copper and other metals in floodplain waters adjacent to the ARD patches, is an increasing cause for concern. The levee ARD occurs primarily during periods of low river levels when levees are exposed and oxidizing, after which localized rainfall leaches metals from the levees resulting in elevated metals in adjacent surface waters (A. Storey, personal communication). OTML 2005 monitoring data show that metals released include copper, aluminum, cadmium, manganese, magnesium, nickel, zinc, and lead. In situ toxicity tests are proposed to determine the ecotoxicity of levee ARD. These will involve sampling of resident flora/fauna to assess community effects and field exposure trials to determine responses in selected taxa deployed in 96-h exposure experiments. Given that food web studies showed that periphyton provided an important carbon source supporting aquatic food chains in the Fly River system, further work is underway to determine the effect of both increased bioavailable copper concentrations and increased turbidity on periphyton communities in the Fly River. Periphyton samples collected from both the Middle Fly and Kiunga are currently being examined for species diversity and abundance, and the sensitivity/tolerance to copper and turbidity of isolated species is being compared. Further research is aimed at developing an algal model to predict algal biomass responses to contaminants in field populations. There is also a lack of basic ecotoxicological data on the effects of copper on primary consumers (algal grazers) such as small invertebrates and zooplankton, planktivorous fish, and other key aquatic organisms in the Fly River system. This is another important knowledge gap currently being addressed through the use of in situ bioassays with field-collected invertebrates. Alternative, more robust methods to ASV to measure labile copper have recently been developed in our laboratory and applied to Fly River water samples ( Apte et al., 2005 ; Bowles et al., 2006 ). The water sample is passed rapidly through a small Chelex resin column, which captures only ionic metal plus metal complexes that are able to dissociate within the short contact time with the resin. Similar to ASV, colloidally bound metals and metals bound in strong complexes do not interact with the resin. 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