toxicity assessment of leachates binet et al - · pdf filetoxicity assessment of leachates...

7

AUSTRALASIAN JOURNAL OF ECOTOXICOLOGY

*Author for correspondence, email: [email protected]

Binet et alToxicity assessment of leachates

Vol. 9, pp. 7-18, 2003

TOXICITY ASSESSMENT OF LEACHATES FROM HOMEBUSH BAY LANDFILLS

MT Binet*1, MA Adams1, JL Stauber1, CK King1, CJ Doyle2, RP Lim2 and E Laginestra3

1 CSIRO Energy Technology, Private Mail Bag 7, Bangor, NSW 2234, Australia.2 University of Technology, Sydney, Westbourne St, Gore Hill, NSW 2065, Australia.3 Sydney Olympic Park Authority, Locked Bag 3, Homebush Bay, NSW 2127, Australia.

ABSTRACTAs part of a pilot ecotoxicological monitoring project, the toxicities of leachates (4 to 29‰ salinity) from landfill mounds in Olympic Park,Homebush Bay, NSW, were assessed using the alga Nitzschia closterium and the fish larva Macquaria novemaculeata. The leachatescontained a mixture of chemicals including metals, ammonia, phenols, BTEX, PAHs and traces of other chemicals, typical of landfills.Both algal growth and fish imbalance tests were sensitive to the leachates, with EC50 values generally less than 10% of the undilutedleachates. There was little temporal variability in leachate toxicity over the three sampling periods for most sites. Spatial variability inleachate toxicity within landfill mounds was examined at two sites. At one of the two sites, there were significant differences (p < 0.05) inthe toxicity to algal growth of leachates from within the mounds. Toxicity of the leachates from all sites and within landfill mounds wascorrelated with total ammonia in the leachates. Toxicity Identification Evaluation (TIE) with the alga confirmed that ammonia was themajor toxicant in a representative leachate.

Key words: toxicity, landfill leachate, ammonia, algae, fish.

INTRODUCTIONUnauthorised and uncontrolled dumping of over nine million cubicmetres of domestic, commercial and industrial wastes aroundHomebush Bay was commonplace from the 1950s to the 1980s.This resulted in extensive contamination of surface andgroundwaters, soils and sediments around the bay (Hayward 1998).Following extensive investigations undertaken by the HomebushBay Development Corporation (later the Olympic Co-OrdinationAuthority [OCA]), remediation of the site commenced in 1994, inpartial fulfilment of a commitment to a “green” Olympic Gamesfor September, 2000. About 20% of the 760 hectare HomebushBay site required remediation to restore its natural habitatcomprising salt marsh and mangrove wetlands, grasslands and forest(OCA 2000).

The landfill remediation strategy focussed on the consolidation andisolation of wastes and on the management of surface waters andleachates emanating from them. Waste was relocated to four largecontainment mounds, which have been capped and re-landscaped.An extensive drain network, designed to prevent leachate frominfiltrating waterways and wetlands, collects water emanating fromthe mounds, together with leachate from five smaller sites. Theleachate is treated on site before being discharged into the sewer.Although regular chemical monitoring of the leachate and surfacewaters is conducted, little is known about the potential toxicity ofthe leachate to aquatic organisms if it were to infiltrate surface andground waters. The chemical compounds or mixtures in the leachatethat may be responsible for causing toxicity to aquatic organismsare also unknown.

Leachates from landfills are generally highly contaminated withammonia, metals and organics such as halogenated hydrocarbons(Clement and Merlin 1995; Bras et al. 2000). Landfill leachateusually also contains high concentrations of inorganic salts includingsodium chloride, carbonate and sulfate (Fatta et al. 1998). Thecomposition of the leachate, however, is influenced by refusecomposition, age, site conditions and runoff (Li and Zhao 1999).

For example, leachates from young landfills with large deposits ofdomestic waste often have high concentrations of ammonia andorganics (Irene and Lo 1996). In contrast, leachates from old landfillsor from industrial waste dumpsites have been reported as havinglower ammonia and organic nitrogen concentrations as the amountof nitrogen available for leaching decreases over time (Burton andWatson-Craik 1998). Ammonia is thought to be produced in landfillsby microbially-mediated protein deamination and by nitratereduction during anaerobic decomposition of the refuse (Burtonand Watson-Craik 1998).

The aim of this study was to assess the toxicity to aquatic biota ofleachates collected from landfill mounds at nine sites in OlympicPark. Each landfill has a different history of contamination andcontains different wastes (OCA 2000). This research formed partof an initial pilot project designed to provide baseline data on thetoxicity of contaminants in leachates to enable improved leachatemanagement and effective recommendations for long-termmonitoring. Temporal and spatial variability in leachate toxicitywas determined over a three-month period (June to August, 2000)using aquatic bioassays with bacteria, microalgae, sea urchins andfish, together with two mammalian cell bioassays (human liver cellsand beef heart mitochondria). Only the microalgae and larval fishtoxicity tests are reported here. The specific objectives were to: 1)establish a baseline of the toxicity of the leachates between andwithin landfill mounds; 2) compare chemical analyses of theleachates with ecotoxicity testing results to determine possiblecauses of leachate toxicity; and 3) enable selection of the mostsensitive and reproducible toxicity tests for long-term leachatemonitoring.

Australian bass (Macquaria novemaculeata) is an importantestuarine-spawning recreational fish species (Harris 1986) and isalso one of the few estuarine fish species that has successfully beenused to assess the toxicity of complex mixtures in Australia (Gulecand Holdway 2000). This sub-lethal acute bioassay determinesimbalance of bass larvae in various concentrations of leachate over96 h in a static test.

8


The microalga Nitzschia closterium is a diatom commonly foundin Australian coastal waters. This species is very sensitive to bothorganic and inorganic compounds (Stauber 1995). The chronicbioassay measures the decrease in cell yield of the alga over 72 h invarious dilutions of leachate compared to salinity-matched controls(Stauber et al. 1994). This bioassay, which is highly reproduciblecompared to higher organism tests, is widely used for monitoringindustrial discharges throughout Australia.

MATERIALS AND METHODS

Leachate collectionTemporal variation studyComposite leachate samples were collected from nine sites inOlympic Park (Figure 1): Clay Pit (CP), Sydney InternationalAquatic Centre Carpark (SIAC), Golf Driving Range (GDR),Archery Park (AP), Haslams Creek South (HCS), Haslams CreekNorth (HCN), Newington/Woo-la-ra (NEW), Wilson Park (WP),and Auburn/Hardy’s Tip (AHT). As all nine sites could not be testedsimultaneously, samples were collected weekly from three of thenine sites over a period of nine weeks, from June to August, 2000.Each site was therefore sampled three times, with the exception ofHCS, which was sampled twice. Samples were held at 4°C beforeuse in the toxicity tests.

Mound spatial variation studySpatial variation in leachate toxicity within mounds, was assessedat two sites (HCS and NEW). Three different manholes or pumppits at each mound were sampled twice over a four week periodfrom July to August, 2000.

Chemical analysesAll leachate samples were analysed for both total and dissolvedmetals and metalloids, together with non-metallic inorganics at theCentre for Advanced Analytical Chemistry (CAAC), a NationalAssociation of Testing Authorities (NATA)-accredited laboratory.The elements boron, barium, chloride, chromium, copper, iron,manganese, nickel and zinc were measured using inductivelycoupled plasma atomic emission spectrometry (ICP-AES); silver,cadmium, cobalt and lead were measured using inductively coupledplasma mass spectrometry (ICP-MS); arsenic and selenium weremeasured using hydride generation atomic fluorescencespectrometry (HGAFS); mercury was measured using cold vapouratomic fluorescence spectrometry (CVAFS); and total cyanide wasmeasured by a continuous flow pyridine-barbituric acid colorimetricmethod. Ammonia was measured using a flow injection method(APHA, 4500-NH

3) at the NATA-accredited laboratories of

Australian Laboratory Services (ALS).

Organic chemical analyses of all leachate samples were carried outby ALS. Organics analysed included polycyclic aromatichydrocarbons (PAHs), total petroleum hydrocarbons (TPH), andBTEX (benzene, toluene, ethylbenzene and xylene) using gaschromatography/mass spectrometry (GS/MS) and phenols (APHA1998). Dioxins (full congener polychlorinated dibenzo-p-dioxins/dibenzofurans (PCDD/Fs) including 2,3,7,8-tetra-chlorodibenzo-p-dioxin (TCDD) and octa-chlorodibenzo-p-dioxin (OCDD)) weremeasured using isotope dilution mass spectrometry by AgriQuality(Wellington, New Zealand), accredited in New Zealand.

Toxicity testsFish larval imbalance testThis sub-lethal acute bioassay determines immobilisation ofAustralian bass larvae, Macquaria novemaculeata (Steindachner)in various concentrations of leachate over 96 h in a static test. Dueto the lack of availability of fish larvae during the test period, toxicitytests were only carried out on one or two samples per site.

Fish larvae (<24 h old) of approximately 5 to 10 mm in lengthwere purchased from Searle Aquaculture, NSW. Seawater wascollected from Rose Bay, Sydney, stored in 15 000 L capacity epoxy-lined concrete tanks and aerated. Prior to use in the tests, the seawaterwas filtered twice (50 and 5 µm) and then diluted to a salinity ofapproximately 25‰ with reverse osmosis water. Twelve leachatesamples were tested over a 4-week period. Leachate samples wereadjusted to a salinity of 25‰ using hyper-saline brine. Sampleswere then diluted to the appropriate concentrations with filteredseawater (25‰). Toxicity tests were commenced within 48 h ofsample receipt.

Five concentrations of each sample, as well as a natural seawatercontrol (25‰) and a hypersaline brine control (diluted to 25‰

Toxicity assessment of leachates Binet et al

Vol. 9, pp. 7-18, 2003

Figure 1. Homebush Bay remediated landfill systems and leachatecollections sites: Clay Pit (CP), Sydney International Aquatic Centre Carpark(SIAC), Golf Driving Range (GDR), Archery Park (AP), Haslams Creek South(HCS), Haslams Creek North (HCN), Newington/Woo-la-ra (NEW), WilsonPark (WP), and Auburn/Hardy’s Tip (AHT). (Taken from OCA 1999)

9


using reverse-osmosis treated water) were prepared. Five hundredmillilitres of each test solution was placed into 1 L glass containers.Four replicates for each concentration and for both controls wereprepared. Five fish larvae were randomly selected and placed intoeach container. Tests were maintained at a temperature of 19 ± 1°Con a 16 h light:8 h dark cycle in environmental chambers. All testcontainers were aerated at a rate of approximately 60 bubbles/minduring the tests. Temperature, salinity, pH and dissolved oxygenwere measured at the start of the test (0 h) and at test termination(96 h). The number of imbalanced larvae (ie. fish unable to maintaintheir balance and swimming ability) was determined at 24-hintervals. Imbalanced fish were removed immediately andanaesthetized in MS222. The test was terminated at 96 h.

The test was considered acceptable if the proportion of balancedfish in controls was ≥90%. Physico-chemical conditions measuredat 0 and 96 h were required to meet the recommended values forfish toxicity tests (OECD 1987). Copper was used as a positivecontrol and was tested in duplicate at six nominal concentrations(31 to 1000 µg/L) to determine if test sensitivity varied with eachbatch of larvae.

Microalgal growth inhibition testThis test determines the inhibition of growth (cell yield) of themarine alga Nitzschia closterium over 72 h. The test is based on theOECD Test Guideline 201 (1984) and the protocol of Stauber et al.(1994), with modifications to test volume (6 mL minivials ratherthan 50 mL flask bioassays).

The unicellular alga N. closterium (Ehrenberg) W. Smith (StrainCS 5) was originally isolated from Port Hacking, NSW. The diatomwas cultured in f medium (Guillard and Ryther 1962) with the ironand trace element concentrations halved. The culture wasmaintained on a 12 h light:12 h dark cycle (Philips TL 40 Wfluorescent daylight, 60 µmol photons/s/m2 at 21°C. Cells in logphase growth were used in the algal bioassays according to thestandard protocol (Stauber et al. 1994). The inoculum was washedand centrifuged three times to remove the culture medium and toconcentrate the algae prior to their use in the bioassays.

Algal tests were commenced within 24 h of sample receipt. Onarrival, salinity, pH, conductivity and dissolved oxygen weremeasured. The samples were adjusted to a salinity of 20‰ by theaddition of dry AR grade GP-2 artificial sea salts (USEPA 1994).Samples were also pH adjusted to 8.1 ± 0.1 using 0.1M and 1MNaOH or HCl. The standard protocol for this species requires filteredsample to be tested; however, in order to determine whether thefiltration process affected the toxicity of the sample, oneconcentration of unfiltered leachate was also tested. The remainingsalinity-adjusted samples were filtered through an acid-washed(10% HNO

3) 0.45 µm membrane filter.

Leachate was tested at a salinity of 20‰ to reflect brackish water/estuarine conditions encountered at the site. Although N. closteriumtests are normally carried out at 34‰, previous studies by CSIROhave shown that there is no effect on the growth rate of N. closteriumin salinities from 15-35‰ (unpublished data). Two natural seawatercontrols (at 20‰ and 34‰), as well as artificial sea salt controls(20‰), were tested with each bioassay for quality assurancepurposes. Two copper tests (2 to 80 µg Cu/L) were run at 20‰salinity, and one test was run at 34‰ (5 to 80 µg Cu/L) to assess

the sensitivity of the alga at the test salinity and to determine theEC50 for copper. Copper was then used as a positive control andwas tested at one nominal concentration (10 µg/L) in each bioassayto ensure that the algae were responding to a known referencetoxicant in a reproducible way.

Each of the controls and the reference toxicant were prepared intriplicate. Natural seawater controls (20‰ and 34‰) were preparedusing 0.45 µm filtered seawater collected from Port Hacking, NSWand diluted with Milli-Q water where appropriate. Artificial seasalt controls (20‰) were prepared by adding GP-2 sea salts to Milli-Q water and then filtering (0.45 µm) through an acid-washed (10%HNO

3) polycarbonate filter unit.

Five leachate concentrations, each in triplicate, were prepared bydiluting the leachate with filtered seawater (20‰). Six millilitres ofeach concentration was dispensed into 20 mL glass scintillationvials that had been silanised (Coatasil, Ajax Chemicals). To eachvial, 0.06 mL of 25 mM sodium nitrate and 0.06 mL of 1.6 mMpotassium dihydrogen phosphate were added as nutrients. Each vialwas inoculated with 2.6 to 4.0 x 104 cells/mL of a prewashed algalsuspension. Vials were incubated at 21°C on a 12:12 h light/darkcycle (Philips LTD 36 W fluorescent daylight) at 150 µmol photons/s/m2. Additional vials at each concentration were prepared formonitoring pH at the beginning and at the end of the test.

Cell density in each treatment was determined after 72 h using eithera Coulter Multisizer II Particle Analyser with 70 µm aperture orusing a hemocytometer and phase contrast microscope (300 xmagnification). Tests were considered acceptable if the final celldensities in the controls were greater than 2x105 cells/mL, with <20% variability. Algal growth was expressed as cell yield. Percentagegrowth inhibition in each treatment was calculated using thefollowing equation:

where:I is the percentage inhibition of algal growth for each test-

concentration replicate;Rc is the mean cell yield for the control; andR is the cell yield for each test-concentration replicate.

Statistical analysesData for M. novemaculeata were arcsine transformed. NOEC (noobservable effect concentration) and LOEC (lowest observable effectconcentration) values were determined using the non-parametricSteels Many-One Rank Test at p ≤ 0.05. EC50 values weredetermined using the Trimmed Spearman-Karber method (Hamiltonet al. 1977).

EC50 values for N. closterium were calculated using LinearInterpolation (ToxCalc Version 5.0.23, Tidepool Software). Aftertesting the data for normality (Shapiro-Wilk’s Test) and homogeneityof variance (F-Test), Dunnett’s Multiple Comparison Test (p ≤ 0.05)was used to determine LOEC and NOEC values. The toxicity ofunfiltered samples was compared to that of filtered samples, byexpressing each as a percentage of their respective control, withanalysis by either homeoscedastic or heteroscedastic t-tests,depending on data normality and homogeneity of variance.


Vol. 9, pp. 7-18, 2003

I = _____ x 100Rc - R

Rc

10


Using the statistical package Statistica, analysis of variance(ANOVA) and Student Neuman Keuls (SNK) tests on EC50estimates were used to test for temporal differences in the toxicityof leachates to both M. novemaculeata and N. closterium and forspatial differences in the toxicity of leachates to N. closterium withinmounds. Cochran’s C-test was used prior to ANOVA to test forhomogeneity of variances (p ≤ 0.05) and EC50 values used in theanalysis were first arcsine transformed to normalise the proportionaldata (Zar 1984).

Toxicity Identification Evaluation (TIE)During the second round of sampling, it was hypothesised thatelevated ammonia concentrations in the leachates were contributingto the observed toxicity to the algae and fish. To test this, a partialTIE (correlation and deletion) was performed on a toxic leachatesample from CP containing 355 mg NH

3-N/L.

Ammonia correlationThe correlation approach involves regressing observed toxicityagainst expected toxicity due to measured concentrations of asuspect toxicant (Burkhard and Ankley 1989). The toxicity of thesuspect toxicant, ammonia (as NH

4Cl) to M. novemaculeata and

N. closterium was determined, using concentrations ranging from0.65 to 50 mg NH

3-N/L. Predicted toxicity (in Toxic Units) of the

leachates (based on ammonia concentrations alone) was determinedby dividing the ammonia concentration in the sample by theammonia EC50 value. The observed toxicity of the leachate intoxic units (TU = 100/EC50) was then compared with the predictedtoxicity.

Toxic unit ratios (TUR = predicted TU/observed TU) were alsocalculated for each leachate sample. A TUR <1 indicated that theleachate toxicity was underestimated (more toxic than predicted)

based on total ammonia. A TUR >1 indicated that the leachatetoxicity was overestimated (less toxic than predicted) based ontotal ammonia.

Ammonia deletionLeachate toxicity to N. closterium was determined before and afterammonia removal. Leachate was diluted to 8% so that theconcentration of ammonia in the leachate was low enough to beremoved by aeration within 48 h. The sample (100 mL in a silanisedglass flask) was adjusted to pH 10 with 1M NaOH, and ammoniawas then blown off the solution by bubbling with air for 42 h. Thevolume of leachate remaining was measured and adjusted withMilli-Q water to the original volume. The pH of the leachate wasreadjusted to 8.1 with 1M HCl. The leachate was tested for toxicityat a range of concentrations (0.5 - 8%) both before and afterammonia removal. The total ammonia concentration wasdetermined before and after aeration, using the MerckSpectroquant® kit (Cat. No. 14752). Seawater and artificial seawaterblanks were prepared using the same pH adjustment/aerationprocedure as for the leachate.

During the aeration procedure, it is possible that any surfactantspresent in the leachate may also be removed by adsorbing onto thesurface of the glass flask through sublation (Ankley et al. 1990).To assess this, the aeration procedure included two extra flasks ofleachate aerated for 42 h. Leachate was carefully removed fromeach flask and the glass rinsed with either seawater or methanol toremove any adsorbed material. These washings were then dilutedin seawater (20‰) and tested for toxicity to algal growth. If areduction in toxicity following aeration was due to the loss ofsurfactants, then this procedure would be expected to recovertoxicity from the glassware (Ankley and Burkhard 1992).


Vol. 9, pp. 7-18, 2003

11


To compare the toxicity of the CP leachate (8%) to N. closteriumfollowing the deletion procedure (TU

deletion) to the toxicity of the

leachate in the baseline test (TUbaseline

), the percentage toxicityreduction due to the deletion procedure (TR

deletion) was calculated

using an equation derived from Van Sprang and Janssen (1997):

%TRdeletion

= (1 – TUdeletion

/TUbaseline

)*100

A positive TRdeletion

value indicates a decrease in toxicity while anegative value indicates an increase in toxicity compared with thebaseline value.

RESULTS

Leachate toxicity at nine sitesFish larval imbalance testThe mean percentage of control fish larvae that remained balancedover 96 h was >90%, indicating test acceptability. Larval sensitivityto copper was consistent over the duration of the study, with theEC50 ranging from 144 to 203 µg Cu/L with a mean of 178 ± 25µg Cu/L. No historical copper data for this laboratory and test wereavailable for comparison, as the copper reference toxicant test wasnot previously used due to animal ethics constraints on fish numbersallowed for toxicity testing. The measured physico-chemicalparameters of the test solutions were maintained within acceptablelimits for each test (OECD 1987). Salinity ranged from 24 to 31‰,temperature ranged from 18 to 21°C, pH ranged from 7.4 to 8.9and dissolved oxygen ranged from 73 to 105% saturation forall tests.

Leachates were toxic to the fish larvae, with average EC50 valuesless than 40% leachate for all sites (Table 1, Figure 2). There wassome temporal variability in leachate from two of the three sitestested. The WP and NEW leachates increased in toxicity from July(EC50 values of 21% and 8.3%) to August (EC50 values of 2.7%and 2.3%).

Due to the limited data set, statistically-significant differencesbetween the toxicity of leachates from different sites could not bedetermined. However, the most toxic sites (from most toxic to least

Figure 2. Toxicity of leachates to Nitzschia closterium and Macquarianovemaculeata. Mean EC50 values (± SE) for tests on leachates throughoutthe study are shown. Standard errors marked by an asterisk (*), could notbe calculated for some sites. Site abbreviations as listed for Figure 1.

toxic) appeared to be AHT, CP, WP (August sample), NEW, AP,HCS and GDR, all with average EC50 values ≤ 13% leachate.Minimal toxicity was observed for leachates from HCN and SIACwith EC50 values of 27% and 39% respectively.

Microalgal growth inhibition testNatural seawater controls (34‰) gave cell yields of 56 ± 11 x 104

cells/mL throughout the study, indicating test acceptability. Naturalseawater controls (20‰) and artificial seawater controls (20‰)were similar to controls at 34‰ salinity, with 38 ± 11 x 104 cells/mL and 41 ± 7.1 x 104 cells/mL respectively, indicating that testingat a salinity of 20‰ did not affect the algal response. On twooccasions (7th and 28th of August, 2000), the natural seawatercontrols (20‰) had lower than acceptable growth rates. Therefore,artificial seawater controls (20‰) were used for cell yieldcalculations and statistical comparisons on these two occasions.

The pH change in all controls over the 72-h test duration was <1pH unit, indicating test acceptability. The reference toxicant copper(10 µg/L) gave 47 ± 24% growth inhibition, indicating an expectedand consistent response to a known toxicant over the duration ofthe study. There was no significant difference (p ≤ 0.05) betweenthe toxicity of copper to N. closterium cell yield at 20 or 34‰salinity. Copper calibration curves in natural seawater at 20‰ and34‰ showed that copper was toxic to N. closterium with EC50values of 9 (5.3-14) and 14 (6.1-30) µg Cu/L at 20‰, and 13(9-19) µg Cu/L at 34‰.

Filtered leachates were generally toxic to the microalga N.closterium, with EC50 values ranging from 1.7% for CP to > 100%for HCN (Table 2, Figure 2). There was little temporal variabilityin leachate toxicity for most sites, with similar EC50 values withina site over the nine-week winter sampling period. However, fortwo of the nine sites, HCS and HCN, toxicity did vary over time,making the variances heterogeneous. Statistical analyses weretherefore carried out after removing these two sites from the dataset. Significant differences between sites (p ≤ 0.05) were detected.The order of toxicity (from most toxic to least toxic) was:

(CP=AHT, AHT=NEW) > SIAC > (AP=GDR=WP)

where “=” is not significantly different and “>” is p ≤ 0.05.

The concentration of unfiltered leachate tested with each bioassaydepended on the expected toxicity of the leachate, and ranged from1.7% for the most toxic sites (eg. CP) to 100% for the less toxicsites (eg. HCN). Unfiltered leachates from SIAC, AP, HCS, HCN,and AHT were of similar toxicity to the alga as the filtered leachates.Unfiltered leachate from the GDR on one occasion (3 July, 2000)was more toxic (p ≤ 0.05) than its corresponding filtered leachate,with 100% growth inhibition after 72-h exposure. Unfilteredleachates were significantly less toxic (p ≤ 0.05) than filteredleachates for both NEW and WP on the 17 July, 2000, and for CPon 3 July, 2000.

Mound spatial variation studyThe toxicity of the leachates collected from HCS to algal growthvaried substantially between the different manholes (Figure 3).However, toxicity of the leachate from each manhole was similarfor the two leachate samples over time.

Within HCS, there was no significant difference (p > 0.05) betweenthe toxicity of filtered and unfiltered leachates. However, there were


Vol. 9, pp. 7-18, 2003

12


some differences in the relative toxicity of filtered and unfilteredleachate samples within the mound during August, 2000. Unfilteredleachates from manholes 1 and 4 caused a significant stimulationof cell yield (p ≤ 0.05) that the filtered leachates did not. Algalgrowth in the filtered leachate from manhole 1 was significantlyless (p ≤ 0.05) than the control, while growth in the unfilteredleachate was not significantly different to the control.

There was little variability in the toxicity of leachates from thedifferent pump pits within the NEW mound to algal growth overthe two sampling periods. All samples were highly toxic to thealga, with EC50 values ranging from 3.9 to 6.6%. The toxicity ofunfiltered leachates was significantly less (p ≤ 0.05) than filtered

leachates for two of the three pump pits (PP2 and PP3) within NEWmound.

Chemical analysesChemical analysis of the leachates together with water qualityguideline trigger values for surface waters taken from ANZECCand ARMCANZ (2000) are listed in Table 3. No guidelines existfor groundwaters or leachates so comparisons of leachate chemicalconcentrations with trigger values are only relevant if leachates wereto infiltrate into surface waters. Nevertheless, these comparisonsprovide a useful tool for suggesting potential causes for concernand for directing site management. The leachates contained highconcentrations of inorganic salts (such as chloride) as is typical of


Vol. 9, pp. 7-18, 2003

13



Vol. 9, pp. 7-18, 2003

14


landfills with mixed municipal and industrial waste. Total ammoniaconcentrations were high for most of the sites. Only on one occasion(at HCN) was the ammonia level lower than the trigger value of0.91 mg/L for surface waters. Un-ionised ammonia (NH

3) was

approximately 4% to 11% over the pH range of the leachates.

Cyanide concentrations in the leachates ranged from <20 µg/L inAHT leachate to 440 µg/L in NEW leachate. Metals includingarsenic, cobalt, copper, zinc, and to a lesser extent chromium, werepresent in the leachates at concentrations high enough to potentiallycause toxic effects to aquatic organisms.

Leachates from NEW and WP contained the widest range of organiccontaminants at elevated concentrations. Both mounds contain tarsludge waste, which leach BTEX, PAHs and phenols. Dioxins areassociated with fly-ash type landfills and were detectable in traceamounts in most leachates, with the highest concentrations at HCN,historically a major site for fly-ash disposal.

TIEAmmonia correlationFor the fish larvae, there was a significant correlation (p ≤ 0.05; R2

= 0.46) between the observed and predicted toxicity based on totalammonia concentrations in the leachates (Figure 4a). Ammoniawas the major toxicant in leachate from SIAC with a TU ratio of1.2. All the other leachates were more toxic to fish larvae than thatpredicted on the basis of the total ammonia concentration alone.

There was a strong and highly significant correlation (p ≤ 0.01; R2

= 0.92) between the predicted and observed toxicity of leachatesfrom each site and within the two mounds, NEW and HCS to thealga N. closterium (Figure 4b, c). TURs for samples from CP, AHT(on two occasions), GDR and NEW were about 1 (Figure 5),suggesting that ammonia was the major toxicant in these leachates.The toxicity of leachate from SIAC on two occasions was less thanthat predicted from the ammonia concentrations in the leachate (ie.TUR >1), suggesting that ammonia toxicity may have beenameliorated by some other compound(s) in the leachate in thesesamples. Leachate samples from AP, HCN, HCS, AHT (third sampleonly), SIAC (third sample only) and WP were more toxic than thatpredicted from the ammonia concentration alone (ie. TUR <1). Inthese samples, other unidentified compounds may also have beencontributing to the observed toxicity.

Figure 3. Toxicity of leachates from manholes within Haslams Creek South(HCS) and Newington (NEW) mounds to Nitzschia closterium. PP2, PP3 =Pump Pits 2 and 3 respectively; NL4 = code number for another samplingsite within Newington.

Ammonia deletionLeachate from CP (containing largely putrescible waste) was oneof the most toxic to algal growth and also had consistently hightotal ammonia concentrations (Table 3). A manipulation on aleachate sample from CP (3 July, 2000) was therefore carried outto confirm that ammonia was the major toxicant at this site. Beforeaeration, the total ammonia concentration in the leachate was 425mg NH

3-N/L and in the 8% dilution was 34 mg NH

3-N/L. Following

overnight aeration and subsequent pH adjustment to 8.1, totalmeasured ammonia was reduced to 0.5 mg/L, which was belowthe LOEC of 7.2 mg/L for toxicity to N. closterium.

Toxicity of the leachate to algal growth was reduced followingammonia removal (Figure 6), with the EC50 value increasing from1.8% to 6.1%. This represents a 70% reduction in toxicity due tothe deletion procedure (TR

deletion). However, inhibition of algal

growth was still observed at the highest concentration tested. It islikely that a precipitate, formed during back adjustment of the pHfrom 10 to 8.1, was responsible for the remaining toxicity. Similartoxicity due to precipitate in the highest test concentration only hasbeen observed in algal TIEs with effluents (Stauber and Adams,unpublished).

The methanol rinse was not toxic to algal growth, with mean cellyields the same as that of the controls. The seawater rinse was alsonot toxic, but did stimulate algal growth by 37%. These resultssuggest that the loss of leachate toxicity after aeration at high pHwas due solely to removal of ammonia and not due to losses oforganics such as surfactants to the silanised glass flasks.

Figure 4. Comparison of predicted toxicities with observed toxicities for(a) Macquaria novemaculeata at all sites, (b) Nitzschia closterium at all sitesand (c) Nitzschia closterium within Haslams Creek South (HCS) andNewington (NEW) mounds. Each line on the graphs represents significantlinear regressions (p ≤ 0.05).


Vol. 9, pp. 7-18, 2003

15


Figure 5. Toxic unit ratios for all Nitzschia closterium tests. A toxic unit ratio <1 indicates the leachate is more toxic than predicted based ontotal ammonia. A toxic unit ratio>1 indicates the leachate is less toxic than predicted based on total ammonia. Site abbreviations are as listedfor Figure 1.

Figure 6. Toxicity of Clay Pit leachate to Nitzschia closterium with and without ammonia removal. Each line represents a concentration-response curve. Values are represented as mean cell yield ± 1 SD.


Vol. 9, pp. 7-18, 2003

16


DISCUSSIONThe leachates from Sydney Olympic Park were toxic to both larvaeof the fish Macquaria novemaculeata and the alga Nitzschiaclosterium suggesting that these leachates may have potentialadverse effects if they were to enter surrounding wetlands andsurface waters. It appears that each species was responding todifferent contaminants within the typically complex mixture ofchemicals that comprised the leachates. There was little temporalvariability in the toxicity of the leachates over the nine-week wintersampling period with the exception of HCS and HCN. The lowwinter rainfall may have contributed to the consistency in theleachate toxicity over this period for most sites.

Considerable spatial variation in the toxicity of leachate from HCSwas found, as is typical of a landfill in various phases ofdecomposition. Waste decomposition in a given landfill occurs infive biological phases from short-term aerobic to aerobic acidfermentation, intermediate anaerobiosis, methanogenic fermentationand secondary aerobic processes (Li and Zhao 1999). Although thefirst phase is usually completed within a month (Irene and Lo 1996),the time it takes for complete waste decomposition to occur willdepend largely on the size of the landfill, environmental factorssuch as rainfall and the fill material. Even within a landfill, such asHCS, wastes will simultaneously be at a variety of stages ofdecomposition (Robinson et al. 1987). This is likely to result invariable leachate characteristics over time and space.

Ammonia concentrations in leachate from various manholes at HCSalso differed and correlated with leachate toxicity to N. closterium.Ammonia concentrations also vary in different phases of landfilldecomposition, depending on the level of microbiological activity,as it is both a toxic by-product of, and essential nutrient for, thebacteria responsible for the degradation processes (Burton andWatson-Craik 1998). In contrast, the ammonia concentrations andtoxicity of NEW mound leachates from each pump pit were similar,suggesting that decomposition stage and rate were similarthroughout the landfill.

In common with other landfills (Clement and Merlin 1995; Ernst etal. 1994; Wong 1989), ammonia was the major cause of toxicity ofthe leachates, especially to the microalga N. closterium. This issupported by the fact that: (1) leachates from landfills at OlympicPark were very high in ammonia, containing from 2 to 1000 timesthe NOEC for ammonia toxicity to N. closterium; (2) toxicity wascorrelated with leachate ammonia concentrations, with regression(R2) values greater than 0.9; and (3) leachate toxicity to microalgalgrowth was reduced by 70% when ammonia was removed fromone leachate (CP).

For five of the nine sites (CP, AHT, GDR, HCS and NEW), leachatetoxicity was predictable from the leachate ammonia concentrations.Leachates from one site, SIAC, were less toxic than predicted byammonia concentrations, and it is possible that some othercompound(s) ameliorated ammonia toxicity in these samples.Leachates from three sites (WP, AP and HCN) were more toxicthan predicted by ammonia alone. Contaminants that may havecontributed to the observed toxicity include zinc, cyanide anddioxins; however, a causal relationship would need to be confirmedusing TIE.

Some fish species have been shown to be sensitive to ammonia atthe concentrations present in the leachates (ANZECC ARMCANZ

2000). However, the order of toxicity of the leachates did not directlycorrespond with leachate ammonia concentrations. Other likelycontributors to toxicity include zinc (highest at AHT; 2.2 mg Zn/L), arsenic (highest at CP; 14 µg As/L), phenols (highest at WP)and aromatic and polyaromatic hydrocarbons (very highconcentrations at both NEW and WP).

Despite a limited number of tests, bass larvae appeared to be moresensitive to the unfiltered leachates than the algal growth tests usingfiltered leachates. This was probably due to differences in thesensitivity of the test organisms, rather than differences in leachateused (filtered versus unfiltered) as there was generally littledifference in the toxicity of filtered and unfiltered leachates to algalgrowth.

The toxicity of the same leachate samples to the bacteria Vibriofischeri (Microtox®), sea urchin fertilisation and larval development(Heliocidaris tuberculata) and inhibition of ATP in a sub-mitochondrial particle test (SMP) with beef heart cells, wasdetermined concurrently in a separate project co-ordinated bySydney University. In vitro cytotoxicity, using human liver cells,was also tested at the University of New South Wales on the sameleachate samples at the same time. A comparison of the toxicitytests’ sensitivities for each leachate at each site (based on EC50values) is shown in Figure 7.

No single test was consistently the most sensitive to all leachates.Algal growth, fish larval imbalance and sea urchin fertilisation andlarval development were the most sensitive tests, based on EC50values.

The two surrogate mammalian tests (in vitro cytotoxicity and thesub-mitochondrial particle tests) were consistently less sensitive tothe leachates than the aquatic tests. Of the aquatic tests, Microtox®

was also not sensitive to many of the leachates, with EC50 valuesoften greater than the highest concentration tested. However, forleachate from WP, Microtox® was the most sensitive test used. TheMicrotox® test is well known for its sensitivity to organiccontaminants (Kaiser and Devillers 1994) and leachate from WPwas high in BTEX, PAHs and total phenols compared with theother sites.

A large-scale monitoring program, investigating seasonal variationin toxicity of the leachates, including sampling after rain events, iscurrently in progress. In cases where ammonia cannot explainobserved toxicities, TIEs with sea urchin larval development andN. closterium growth inhibition are being used to identify thetoxicants in the leachates.

CONCLUSIONSThe leachates collected from Olympic Park, Homebush Bay, weretoxic to algal growth and to fish larvae with EC50 values rangingfrom 1.7 to >100% for algal growth and from 1.7 to 39% for larvalfish imbalance. Toxicity was correlated with ammonia for algalgrowth, and using TIE procedures, ammonia was identified as themajor contaminant responsible for toxicity at one of the most toxicsites, CP. Ammonia appeared to be acting in combination with otherchemicals within the complex leachate mixture at four of the ninesites, where leachates were more toxic to algae than predicted basedon measured total ammonia concentrations alone. Similarly, itappears that other chemicals in combination with ammonia wereresponsible for the observed toxicity to fish larvae.


Vol. 9, pp. 7-18, 2003

17


Figure 7. Comparison of the toxicity of leachates from nine sites in Olympic Park for seven toxicity tests (mean ± SE). Standard errors marked by anasterisk (*), could not be calculated for some sites. “>” indicates that the EC50 was greater than the highest concentration tested. Site abbreviations areas listed for Figure 1.

There was little temporal and spatial variability in the toxicity ofleachates from Newington/Woo-la-ra to algal growth. All leachatesfrom this mound were highly toxic, with EC50 values ranging from3.9 to 6.6%. In contrast, the toxicity of leachates collected fromHCS varied substantially spatially between manholes, although therewas little temporal variability at this site. For both sites, toxicity toalgae was highly correlated with the leachate ammoniaconcentrations. Variations in the toxicity of leachates within sitesare probably due to variations in microbiological activity and hencedecomposition rates within landfill mounds.

ACKNOWLEDGEMENTSThe authors would like to thank Donald Cheong and MatthewTiltman for the algal bioassays and Rob Rowland, Rosemary Woodand Brett Warden for the inorganic analyses.

REFERENCESAnkley GT and Burkhard LP. 1992. Identification of surfactants astoxicants in a primary effluent. Environ. Toxicol. Chem. 11,1235-1248.

Ankley GT, Katko A and Arthur JW. 1990. Identification of ammoniaas an important sediment-associated toxicant in the lower Fox Riverand Green Bay, Wisconsin. Environ. Toxicol. Chem. 9, 313-322.

ANZECC / ARMCANZ. 2000. Australian and New ZealandGuidelines for Fresh and Marine Water Quality. Australian and NewZealand Environment Conservation Council and the Agricultureand Resource Management Council of Australia and New Zealand,Canberra, Australia.

APHA. 1998. Standard Methods for the Examination of Water andWastewater. 20th Edition, American Public Health Association,American Water Works Association and Water EnvironmentFederation, Washington DC, USA.

Bras I, Santos L and Alves A. 2000. Monitoring organochlorinepesticides from landfill leachates by gas chromatography-electron-capture detection after solid-phase microextraction. J. Chromatogr.A. 891, 305-311.

Burkhard LP and Ankley GT. 1989. Identifying toxicants: NETAC’stoxicity-based approach. Environ. Sci. Technol. 23(12), 1438-1443.

Burton SAQ and Watson-Craik IA. 1998. Ammonia and nitrogenfluxes in landfill sites: applicability to sustainable landfilling. WasteManag. Res. 16(1), 41-53.

Clement B and Merlin G. 1995. The contribution of ammonia andalkalinity to landfill leachate toxicity to duckweed. Sci. TotalEnviron. 170, 71-79.

Ernst WR, Hennigar P, Doe K, Wade S and Julien G. 1994.Characterization of the chemical constituents and toxicity to aquaticorganisms of a municipal landfill leachate. Water Pollut. Res. J.Canada 29(1), 89-101.

Fatta D, Voscos C, Papadopoulos A and Loizidou M. 1998. Leachatequality of a MSW landfill. J. Environ. Sci. Health A33(5),749-763.

Guillard RRL and Ryther JH. 1962. Studies of marine planktonicdiatoms. I. Cyclotella nana and Detonula confervaceae Hustedt(Cleve) Gran. Can. J. Microbiol. 8, 229-239.


Vol. 9, pp. 7-18, 2003

®

18


Gulec I and Holdway DA. 2000. Toxicity of crude oil and dispersedcrude oil to ghost shrimp Palaemon serenus and larvae of Australianbass Macquaria novemaculeata. Environ. Toxicol. 15(2), 91-98.

Hamilton MA, Russo RC and Thurston RV. 1977. TrimmedSpearman-Karber methods for estimating median lethalconcentrations in toxicity bioassays. Environ. Sci. Technol. 11, 714-719. Correction: Ibid 1978 12:417.

Harris JH. 1986. Reproduction of the Australian bass, Macquarianovemaculeata (Perciformes: Percichthyidae) in the Sydney Basin.Aust. J. Mar. Fresh. Res. 37, 209-235.

Hayward K. 1998. Going for Green. Water Qual. Int. August, 1998,18-23.

Irene M and Lo C. 1996. Characteristics and treatment of leachatesfrom domestic landfills. Environ. Internat. 22(4), 433-442.

Kaiser KLE and Devillers J. 1994. Ecotoxicity of chemicals toPhotobacterium phosphoreum. Vol 2, Gordon and Breach SciencePublishers, France, 879 pp.

Li XZ and Zhao QL. 1999. Inhibition of microbial activity ofactivated sludge by ammonia in leachate. Environ. Internat. 25(8),961-968.

OCA. 1999. Environment Report Summary 1999. Olympic Co-ordination Authority, Sydney, Australia.

OCA. 2000. The Big OCA Clean Up: A Community Guide. OlympicCo-ordination Authority Ecology Programs, Sydney, Australia.

OECD. 1984. Guideline for Testing of Chemicals. Alga growthinhibition test. Test Guideline No. 201. Organisation for EconomicCooperation and Development, Paris, France.

OECD. 1987. Guideline for testing of chemicals. Guideline No.203. Organisation for Economic Cooperation and Development,Paris, France.

Robinson HD, Barr MJ and Last SD. 1987. Leachate collection,treatment and disposal. J. Inst. Wat. Environ. Manag. 6(3),321-332.

Stauber JL. 1995. Toxicity testing using marine and freshwateralgae. Australas. J. Ecotoxicol. 1, 15-24.

Stauber JL, Adams MS, Binet MT, King CK, Lim RP, Doyle CJand Markovic B. 2000. Toxicity Assessment of Leachates fromHomebush Bay Landfills. CSIRO Energy Technology Report ET/IR341R. Sydney, NSW, Australia.

Stauber JL, Tsai J, Vaughan GT, Peterson SM and Brockbank CI.1994. Algae as indicators of toxicity of the effluent from bleachedeucalypt kraft pulp mills. National Pulp Mills Research ProgramTechnical Report No. 3. CSIRO, Canberra, Australia.

USEPA. 1994. Short term methods for estimating the chronictoxicity of effluents and receiving waters to marine and estuarineorganisms. USEPA Report No. EPA-600-4-91-003. United StatesEnvironmental Protection Agency, USA.

Van Sprang PA and Janssen CR. 1997. Identification andconfirmation of ammonia toxicity in contaminated sediments usinga modified toxicity identification evaluation approach. Environ.Toxicol. Chem. 16(12), 2501-2507.

Wong MH. 1989. Toxicity test of landfill leachate usingScaotherodon mossambicus (freshwater fish). Ecotoxicol. Environ.Safety 17, 149-156

Zar JH. 1984. Biostatistical Analysis - 2nd Edition, Prentice-HallInc, New Jersey, USA, 718pp.


Vol. 9, pp. 7-18, 2003

toxicity assessment of leachates binet et al - · pdf filetoxicity assessment of leachates...

Documents