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This page is intentionally blank WATER QUALITY RESEARCH JOURNAL OF CANADA

SCOPE OF THE JOURNAL

The Water Quality Research Journal of Canada publishes peer-reviewed, scholarly articles on the following general subject areas:

• Impact of contaminants on aquatic ecosystems • Contributions of pollutants from the gas and solid phases to aquatic systems • Drinking water, wastewater, and stormwater treatment technologies and strategies • Exotic species, biodiversity, and aquatic species at risk • Groundwater remediation • Assessment of surface and subsurface water quality • Regulations, strategies, and policies related to water pollution • Conservation and protection of aquatic environments • Social science issues in relation to water quality • Review articles on all aspects of water quality and pollution control.

The Water Quality Research Journal of Canada is a quarterly publication. It is meant to be a forum for original research dealing with the aquatic environment. Articles from outside of Canada are welcome provided that they are of interest to the Canadian water quality research community (e.g., they provide a new method of analysis that can be applied in the Canadian context, or the data would be of general interest to the research community).

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Janet McAvella Subscription and Membership Offi ce Canadian Association on Water Quality P.O. Box 5050 Burlington, ON L7R 4A6 Canada E-mail: [email protected] WATER QUALITY RESEARCH JOURNAL OF CANADA

Volume 43, No. 4, 2008

Identifi cation of Sphingomonads on the Basis of Polymerase Chain Reaction Amplifi ed 16S rRNA 249–256 Gene S. Tokajian, M. Farah, and F. Hashwa Relative Body Size Infl uences Breeding Propensity in Fathead Minnows: Implications for 257–264 Ecotoxicology Testing Procedure M.S. Pollock, S. E. Fisher, A.J. Squires, R. J. Pollock, D.P. Chivers, and M.G. Dubé Distribution of 14C-labelled Atrazine, Methoxychlor, Glyphosate, and Bisphenol-A in Goldfi sh 265–274 Studied by Whole-Body Autoradiography (WBARG) C. Rouleau and J. Kohli Exposure to Model-Scale Sewage Treatment Plant Effl uent Affects Circulating Sex Steroids in 275–282 Rainbow Trout J.L. Parrott, M.E. McMaster, S. Verma, and D. Trowbridge Swimming in Sewage: Indicators of Faecal Waste on Fish in the Saint John Harbour, New 283–290 Brunswick H.A. Loomer, K.A. Kidd, T. Vickers, and A. McAslan Concentrations of Endotoxins in Waters Around the Island of Montreal, and Treatment Options 291–303 R. Gehr, S. Parent Uribe, I.F. Da Silva Baptista, and B. Mazer Infl uence of Polymer Selection on Nutrient Phase Separation for Waste Activated Sludge Thickening 305–312 at Bench Scale A.J. MacDonald and O.D. Basu Filtration du bleu de méthylène, du chrome hexavalent et de l’acide éthylène diamine tétracétique 313–320

sur une membrane céramique d’ultrafi ltration à base de ZnAl2O4-TiO2 E.G. Chahid, H. Loukili, S. Tahiri, S.A. Younssi, A. Majouli, et A. Albizane (Filtration of Methylene Blue, Hexavalent Chromium, and Ethylenediaminetetracetic Acid

through an Ultra-Filtration Ceramic ZnAl2O4-TiO2-based Membrane) Philip H. Jones Award: 24th Eastern Canadian Symposium on Water Quality Research v 44th Central Canadian Symposium on Water Quality Research Volume 43 Referees vi–vii

Volume 43 Key Word Index viii–ix

Volume 43 Author Index x

i WATER QUALITY RESEARCH JOURNAL OF CANADA

EDITOR

R. Gehr (McGill University)

EDITORIAL BOARD – ASSOCIATE EDITORS

S.A. Andrews (University of Toronto) F. Archibald (Pulp and Paper Research Institute of Canada) J.-F. Blais (Institut national de la recherche scientifi que) P.-Y. Caux (Environment Canada) S.C. Courtenay (Fisheries and Oceans Canada at the Canadian Rivers Institute/University of New Brunswick) M. Giddings (Health Canada) J. Marsalek (Environment Canada) B. Rabinowitz (CH2M HILL Canada Limited) A. St-Hilaire (INRS-ETE, Université du Québec) E. van Bochove (Agriculture and Agri-Food Canada) P. Vanrolleghem (Université Laval) T. Viraraghavan (University of Regina) H. Zhou (University of Guelph)

MANAGING EDITOR

Janet Jardine Environment Canada Canada Centre for Inland Waters 867 Lakeshore Road, P.O. Box 5050 Burlington, ON L7R 4A6 Telephone: (905) 336-4513. Fax: (905) 336-4420 E-mail: [email protected]

EDITORIAL ASSISTANT

Zakiah Taha (Environment Canada)

PAGE LAYOUT

Lucas Neilson (Graphic Arts Unit, Environment Canada)

The Water Quality Research Journal of Canada (www.wqrjc.ca) receives support from Environment Canada’s National Water Research Institute (www.nwri.ca).

ii CANADIAN ASSOCIATION ON WATER QUALITY ASSOCIATION CANADIENNE SUR LA QUALITÉ DE L’EAU

EXECUTIVE COMMITTEE President IWA WWC 2010

President J.P. Jones, Département de génie chimique, Université de R.L. Droste, Department of Civil Engineering, University of Sherbrooke, Québec Ottawa, Ottawa, Ontario SUBSCRIPTION AND MEMBERSHIP OFFICE Vice Presidents J.A. Nicell, Department of Civil Engineering and Applied J. McAvella, Canada Centre for Inland Waters, Burlington, Mechanics, McGill University, Montréal, Québec Ontario (Eastern) C. Marvin, National Water Research Institute, CORPORATE MEMBERS 2008 Environment Canada, Burlington, Ontario (Central) E.R. Hall, Department of Civil Engineering, University Agilent Technologies of British Columbia, Vancouver, British Columbia Wilmington, Delaware, U.S.A. (Western) AMEC Earth & Environmental Inc. Secretary Thorold, Ontario R.D. Tyagi, INRS-Eau, Université du Québec, Sainte-Foy, Québec Anachemia Science Mississauga, Ontario Treasurer C. Allain, Greater Moncton Sewerage Commission, Avensys Inc. Riverview, New Brunswick Mississauga, Ontario

DIRECTORS Can-Am Instrumental Ltd. Oakville, Ontario Eastern Region G. Gagnon, Centre for Water Resources Studies, Dalhousie Campbell Scientifi c Canada Corp. University, Halifax, Nova Scotia Edmonton, Alberta P.A. Vanrolleghem, Département Génie Civil Pavillon Pouliot, Université Laval, Québec City of Mississauga Mississauga, Ontario Central Region G. Krantzberg, Department of Civil Engineering, McMaster City of Hamilton, Public Health University, Hamilton, Ontario Hamilton, Ontario

Western Region Dr. Jarmo Sallanko J.A. Oleszkiewicz, Department of Civil Engineering, University of Oulu University of Manitoba, Winnipeg, Manitoba Oulu, Finland C. Tiedemann, EPCOR Water Services Inc. Edmonton, Alberta EPCOR Calgary, Alberta PAST PRESIDENT Hoskin Scientifi c Ltd. Y. Comeau, Département des génies civil, géologique et des Burlington, Ontario mines, École Polytechnique, Montréal, Québec Region of Peel Brampton, Ontario

Regional Municipality of Niagara Thorold, Ontario

Regional Municipality of York Newmarket, Ontario

IWA GOVERNING BOARD MEMBER

R.L. Droste, President, CAWQ

Identifi cation of Sphingomonads on the Basis of Polymerase Chain Reaction Amplifi ed 16S rRNA Gene

Sima Tokajian,* Maya Farah, and Fuad Hashwa

Department of Biology, Lebanese American University, Byblos, Lebanon

Sphingomonas is a genus that is basically of environmental origin but can also be associated with health hazards, especially in the hospital environment where there is a great need to properly monitor water sources. The abundance and frequent isolation of derivatives of yellow pigmented colonies from drinking water samples in Lebanon—where an intermittent mode of supply is employed, and which induces frequent biofi lm sloughing—necessitated the establishment of a rapid and feasible assay to screen specifi cally for sphingomonads. In this study, 50 isolates recovered from drinking water with yellow- to orange-pigmented colonies were used to establish a polymerase chain reaction-based (PCR-based) screening assay. The use of sphingomonad specifi c modifi ed primers gave one common band with a size of 320 bp in all presumptive and sequence confi rmed sphingomonads. However, no amplifi cation was observed with Escherichia coli, Staphylococcus aureus, and Pseudomonas aeruginosa. Applying the PCR-based assay described in this paper increased both the effi ciency and the reliability of screening for sphingomonads in water samples, thereby minimizing related risk factors.

Key words: drinking water, sphingomonads, PCR, yellow colonies, biofi lm, intermittent supply

Introduction are known to induce corrosion in copper pipes (Arens et al. 1995; White et al. 1996; Busse et al. 1999). Most Sphingomonads are Gram-negative, chemoheterotrophic, sphingomonads are environmental microorganisms, but nonsporeforming, straight rods, strictly aerobic, and some strains have also been associated with nosocomial characterized by an outer membrane that contains infections. Most of these nosocomial infections originate glycosphingolipids as cell envelope components, from contaminated medical devices (indwelling but lacks lipopolysaccharide (Yabuuchi et al. 1990; catheters, bronchofi beroscopes, and ventilators), White et al. 1996). Colonies are yellow-pigmented or solutions, and water (ultraviolet light irradiated water whitish brown (Takeuchi et al. 1993). Sphingomonas is used in surgery and dental unit water lines) (Yabuuchi widespread in water, soil, sediments, and in association et al. 1990; Barbeau et al. 1996; Lemaitre et al. 1996). with plants (White et al. 1996; Balkwill et al. 1997). Contamination of these medical devices occurs when Large numbers of phenotypically and phylogenetically sphingomonads present in biofi lms in water distribution similar strains of Sphingomonas have been isolated from systems recover from their dormant state upon transfer various environments and described as novel species. The to a more hospitable environment (Mossel et al. 2004). nomenclature of the genus Sphingomonas was revised and Sphingomonas paucimobilis associated with septicaemias the genus was divided into four genera: Sphingomonas, in a haematological unit of a university hospital was Sphingobium, Novosphingobium, and Sphingopyxis directly linked to bacterial colonization of the hospital (Takeuchi et al. 2001). water system (Perola et al. 2002). Furthermore, Hsueh Organisms isolated from a drinking water et al. (1998) revealed that S. paucimobilis was widely distribution network and water storage tanks in Lebanon distributed in hospital environments and could be were previously defi ned to be mainly Gram-negative, involved in causing recurrent infections, in hospitalized pigmented, and as belonging to the α-Proteobacteria patients, such as intravascular catheter-related bacteremia, within the confi nes of Sphingomonadaceae (Tokajian bacteremic biliary tract infection, urinary tract infection, and Hashwa 2004a, 2004b, 2004c; Tokajian et al. 2005). ventilator-associated pneumonia, and wound infection. The high incidence of sphingomonads was consistent The identifi cation of heterotrophic bacteria with their diverse metabolic capabilities, enabling them indigenous to the environment, including sphingomonads, to inhabit a wide range of environments (Pollock and has been based on morphological and physiological Armentrout 1999). The isolation of Sphingomonas from characterization of each organism after isolation (Amy drinking water is not desirable. Members of this group et al. 1992). Commercial kits with different diagnostic are known to be pathogens or opportunistic pathogens, substrates have also been employed either as an alternative produce extracellular polymers, enhance biofi lm or as a complementary approach to the conventional formation and resistance towards disinfection, and some identifi cation methods (Braun-Howland et al. 1993). The use of these commercial kits for the identifi cation of environmental isolates is not very useful because of * Corresponding author: [email protected] their limited databases (Amy et al. 1992; Tokajian and

249 Tokjian et al.

Hashwa 2003; Tokajian et al. 2005). On the other hand, 16S rRNA Gene Amplifi cation identifi cation of bacteria based on rRNA gene sequences has in recent years emerged as a more reliable alternative 16S rDNA gene was amplifi ed using the gene sequence to traditional identifi cation (Böttger 1989; Boye et al. specifi c universal forward 27F (5’-AGA GTT TGA TCC 1999). TGG CTC AG-3’) and reverse primers 1492R (5’-GGT This paper describes the use of a 16S rRNA gene- TAC CTT GTT ACG ACT T-3’) (TiBMolBiol, Germany) based polymerase chain reaction (PCR) assay for the (Lane et al. 1985). 16S rRNA gene amplifi cation with a identifi cation of yellow pigmented colonies, which allows product of 1,500 bp was used as a positive PCR control for the accurate detection of all known sphingomonads ensuring the integrity of all DNA samples used for the isolated on R2A from drinking water. We also compared identifi cation of yellow pigmented colonies. in general terms the capability of the Sphingomonas- The amplifi cation was performed on a 1.5 μl-DNA specifi c PCR screening assay and Biolog biotyping based extract in 20 μl using 1U of AmpliTaq Gold polymerase on individual test results. (Applied Biosystems, U.S.A.), 2.5 mM MgCl2, 1x PCR buffer, 0.4 mM of each deoxynucleoside triphosphate Materials and Methods (dNTP), and 0.25 μM of the forward and reverse primers. DNA amplifi cations were performed on a Perkin Water Samples Elmer GeneAmp 9700 thermal cycler. The cycles used were as follows: 1 cycle at 95ºC for 2 min; 30 cycles of The study was conducted using all forms and derivatives denaturation at 94ºC for 30 s, annealing at 53ºC for 30 of yellow pigmented colonies isolated both from an s and elongation step at 72ºC for 2 min. PCR amplicons intermittent drinking water distribution network were visualized by eithidium bromide staining on a 1% (Tokajian et al. 2005), and polyethylene and cast iron agarose gel. household storage tanks in Lebanon over a period of two years (Tokajian and Hashwa 2004b). Sphingomonas-Specifi c 16S rRNA-Based PCR Assay Isolation and Purifi cation of Yellow Pigmented Colonies The Sphingomonas-specifi c primer set consisted of the modifi ed forward primer Sphingo 108f (5’-GCGT Yellow pigmented colonies were isolated and purifi ed AACGCGTGGGAATCTG-3’) and the reverse primer on R2A agar (Oxoid) (Reasoner and Geldreich 1985). Sphingo 420r (5’-TTACAACCCTAAGGCCTTC-3’) Plates were incubated at 28ºC for seven days. Pure (Leys et al. 2004). The PCR mixture contained 2 μl colonies were kept at -20ºC on glycerol. A total of 50 of DNA, 1 U of AmpliTaq Gold polymerase (Applied derivatives of yellow pigmented colonies, out of around Biosystems, U.S.A.), 20 pmol of the forward and reverse 200, representing all the different morphological entities primers, 10 nmol of each deoxynucleoside triphosphate within this population were chosen and used for biotyping (dNTP), 2.5 mM MgCl2, and 1X PCR buffer in a fi nal and 16S rRNA gene-based studies. volume of 50 μl. All PCR assay runs incorporated a reagent control (without template DNA) and a negative Biotyping PCR control using DNA extracted from Escherichia coli and Staphylococcus aureus. PCR amplifi cation The chosen colonies were identifi ed using the Biolog was comprised of the following three steps: heating at (Biolog, Inc., Hayward, California) microbial 95ºC for 5 min; 50 cycles of denaturation at 95ºC for identifi cation system. The metabolic profi le of each 5 s, annealing at 62ºC for 10 s, and extension at 74ºC organism using the Biolog microplates was compared for 30 s; and a fi nal extension at 74 ºC for 2 min. The automatically by using the MicroLog software with expected PCR amplicon was around 320-bp long and the MicroLog GN database (release 4.01A). Biolog was visualized by ethidium bromide staining on 2.5% identifi cations were reported if the similarity index of the agarose gel. genus or species was 0.5 or greater at 24 h of incubation. A phenogram was generated using the UPGMA algorithm 16S rDNA Sequencing (CLC bio A/S, Denmark) based on the metabolic profi les obtained for each of the tested isolates. The 16S rDNA was amplifi ed in a total volume of 20 μl using the primers SSU-bact-27f (5’- AGA GTT TGA TCM DNA Extraction TGG CTC AG -3’) and SSU-bact-519r (5’- GWA TTA CCG CGG CKG CTG -3’) (Lane 1991). The reaction Chosen colonies were designated as ST-1 through ST- contained 2 μl of 50 ng DNA, 200 μM dNTPs, 0.4 pmol 50. DNA extractions were made by suspending several of each primer, 1x PCR Buffer II (Applied Biosystems), colonies from R2A plates that had grown for a minimum 2.5 mM MgCl2, and 0.1 U of AmpliTaq Gold DNA of 5 days at 28ºC in 1 ml of sterile water. DNA was then polymerase (Applied Biosystems). The thermal cycling released by boiling for 15 min. reaction consisted of an initial denaturation for 12 min

250 16S rRNA Gene-Based Identifi cation of Yellow Colonies at 95°C followed by 30 cycles of denaturation (30 s at 94°C), annealing (30 s at 60°C), and extension (1 min at 72°C), with a single fi nal extension for 10 min at 72°C. The amplicons were sequenced using the ABI Prism BigDye Terminator v3.1 Cycle Sequencing Kit (Applied Biosystems). The sequencing products were purifi ed using the Centri-Sep Columns (Princeton Separations, Adelphia, NJ) and analyzed using the ABI 3130xl Genetic Analyzer (Applied Biosystems) according to the manufacturer’s instructions.

Results

Isolation and Puriﬁ cation of Yellow Pigmented Colonies

Figure 1 shows different forms and derivatives of yellow pigmented colonies representing some of the Fig. 1. Different forms and derivatives of yellow pigmented morphological entities recovered from drinking water colonies representing some of the morphological entities samples collected from water storage tanks and the water recovered on R2A agar from drinking water samples col- distribution network over a period of two years. lected from water storage tanks and the water distribution network. Biotyping Sphingomonas-Speciﬁ c 16S rRNA PCR The metabolic profi les obtained, based on the utilization of 95 carbon sources, from the Biolog microplates were The PCR assay based on the use of a sphingomonad compared with ATCC Sphingomonas species as well as specifi c primer set gave a single band with a size of some other sphingomonads that were previously isolated around 320 bp in all wells, except in the ones containing and identifi ed through 16S rDNA sequencing (Tokajian the reagent control, DNA extracted from Escherichia coli and Hashwa 2003; Tokajian and Hashwa 2004b). and Staphylococcus aureus, and two other presumptive Results obtained revealed that there was considerable sphingomonads (ST-39 and ST-41) (Fig. 4). ST-39 and variation among the isolates. Biotype 1 (7 isolates: ST- ST-41 formed brownish to yellow colonies. 16S rDNA 3, 4, 5, 7, 8, 9, and 10) was found to be most similar sequencing revealed that ST-39 and ST-41 were closely to the ATCC strains Sphingomonas macrogoltabida and related to Mycobacterium sp. and Bacillus sp., respectively. Sphingobium yanoikuyae (Fig. 2). Biotype 2 (2 isolates: Some of the isolates including ST-1, 3, 4, 5, 8, 9, 10, 18, ST-12 and 33) was closely associated with Sphingomonas 23, 24, 27, 35, and 36 showed at least two additional sp. strain SRS7. Biotype 3 (4 isolates: ST-21, 39, 41, and bands with a size of 400 to 500 bp. The former isolates 49) was clustered with Sphingomonas adhaesiva and formed on R2A agar dark yellow to orange colonies Sphingomonas sp. strain JSS-7. Biotype 4 (10 isolates: and were distributed in four different biotypes namely: ST-11, 18, 19, 23, 24, 25, 26, 27, 47, and 50) was found Biotype 1 (ST-3, 4, 5, 8, 9, and 10), Biotype 3 (ST-1), to be closely related to Sphingomonas chungbukensis. Biotype 4 (ST-18, 23, 24 and 27), and Biotype 6 (ST-35 Biotype 5 (9 isolates: ST-34, 35, 36, 37, 38, 40, 42, 43, and 36) (Fig. 2). and 48) and Biotype 6 (3 isolates: ST-15, 20, and 31) did not cluster with any of the used reference ATCC strains. 16S rDNA Sequence Analysis Biotype 7 (3 isolates: ST-28, 32, and 46) was found to be closely related to Sphingomonas sp. strain BN6. ST-39 and ST-41 consensus sequences were analyzed Biotype 8 (ST-30) and Biotype 9 (ST-22) were closely against the nucleotide collection database using the related to Sphingobium sp. strain S10 and Sphingomonas NCBI Blast. ST-39 was found to be most similar to natatoria, respectively. Finally the remaining isolates ST- Mycobacterium sp. and a few soil and Actinobactria sp., 17 (Sphingomonas sp. strain IFO 11), ST-14, and ST-2 while ST-41 was similar to Bacillus sp. and a few marine each represented a separate branch, and the overall sediment bacteria, with maximum identities ranging similarity index was very low (Fig. 2). between 96 and 100%.

16S rDNA Ampliﬁ cation Discussion

The quality of the extracted DNA from all sphingomonads The autochthonous microbial population of drinking was evaluated by the amplifi cation of the 16S rDNA (Fig. water resources, representing an oligotrophic nutrient- 3). All samples except for the reagent control showed a deprived habitat, remains largely uncharacterized clear band at 1500 bp. (Kalmbach et al. 1999). Only a small percentage of

251 Tokjian et al.

Sphingomonas macrogoltabida

Sphingobium yanoikuyae

ST-13 (Sphingomonas sp. strain SRS7) ST-1 (Sphingomonas adhaesiva)

ST-16 (Sphingomonas sp. strain JSS-7)

ST-44 (Sphingomonas chungbukensis)

ST-45 (Sphingomonas sp. strain BN6)

ST-29 (Sphingobium sp. strain S10)

ST-17 (Sphingomonas sp. strain IFO 15917)

ST-6 (Sphingomonas natatoria)

Fig. 2. Clustering of drinking water sphingomonads on the basis of the utilization and hydrolysis of carbon sources (UPGMA algorithm). The positions of relevant reference strains are also shown.

252 16S rRNA Gene-Based Identifi cation of Yellow Colonies

Fig. 3. PCR-mediated amplifi cation of the 1500-bp segment of 16S rRNA gene. DNA samples were analyzed on a 1% agarose gel. Well 1: O’RangeRuler 500 bp DNA ladder (Fermentas); Well 2: negative reagent control; Well 3: Staphylococ- cus aureus positive control; Well 4: Escherichia coli positive control; Lanes 1 through 50 presumptive ST-1 through ST-50 sphingomonads.

Fig. 4. PCR-mediated amplifi cation of 320-bp Sphingomonas-specifi c 16S rRNA gene. DNA samples were analyzed on a 2.5% agarose gel. Well 1: Molecular weight marker VIII (Roche); Well 2: negative reagent control; Well 3: Staphylococcus aureus Gram-positive organisms used as a negative PCR control; Well 4: Escherichia coli Gram-negative organisms used as a negative PCR control; Lanes 1 through 50 presumptive ST-1 through ST-50 sphingomonads.

253 Tokjian et al. these bacteria encountered in water distribution systems pattern distinguished those sequenced isolates from all are identifi able, as many fail to grow on conventional other sequenced sphingomonads, with the same pattern media used for biochemical characterization (Spino being also consistent with isolates ST-1, 3, 8, 23, 24, 1985). Yellow pigmented colonies are common in water 27, 35, and 36. All the isolates that showed additional samples collected both from water storage tanks and bands formed dark yellow to orange colonies. At this distribution networks in Lebanon. The identifi cation of stage it is not clear whether this additional band is HPC (heterotrophic plate count) isolated from drinking due to nonspecifi c amplifi cation or due to the presence water samples, based on the sequencing of the 16S rRNA of more than one rRNA operon (Condon et al. 1995; gene, revealed that some are members of the family Kabadjova et al. 2002). Finally, it is noteworthy that Sphingomonodeceae (Tokajian and Hashwa 2004a; two of the presumptive sphingomonads tested in this Tokajian et al. 2005). study having yellow pigmented colonies (ST-39 and ST- Sørensen et al. (2001) used the Biolog system for 41) did not show any amplifi cation products using the the identifi cation of Sphingomonas sp. (SRS2), and Sphingomonas-specifi c PCR assay. 16S rDNA sequencing Yang et al. (2006) tested the ability S. chlorophenolica revealed that ST-39 and ST-41 were closely related to to utilize (oxidize) various carbon sources, while Mycobacterium sp. and Bacillus sp., respectively. This Pollock (1994) identifi ed isolates secreting gellan- indicates that it is very much likely to have other than related polysaccharides as Sphingomonas. The API sphingomonads appearing as yellow pigmented colonies 20NE and Biotype 100 were also used by Yabuuchi et on R2A agar, which additionally supports the use of a al. (2002) for biochemical characterization of different PCR-based assay to screen for sphingomonads from sphingomonads. All biochemical identifi cation schemes drinking water recovered on R2A agar. including the Biolog (Biolog, Inc., Hayward, California) Growing sphingomonads requires 4 to 5 days on and the API (bioMerieux, Marcy-L’Etoile, France) may R2A agar, and employing the PCR-based assay will not yield ambiguous and misleading results (Tokajian et al. reduce the time needed but will defi nitely increase the 2005). The tests used in those schemes may not lead to testing reliability, since all tested isolates (presumptive reproducible results and the phenotype of a species is sphingomonads, previously sequenced isolates and ATCC not an absolute property but may exhibit remarkable strains) gave one common amplifi cation product (320 bp). variability. The database of phenotypic characteristics is This is especially true when taking into consideration that only limited to common species (Springer et al. 1996; Tang biotyping (which currently is the only available method et al. 1998), not necessarily suitable for environmental for identifying isolates of environmental origin including isolates. However, we tried to base phenotyping on sphingomonads) through the use of commercial kits such individual Biolog test results to overcome database as the Biolog, and even when based on metabolic profi les related constraints, but the low similarity indices (Fig. 2) to overcome database limitations, resulted in very low obtained upon clustering those isolates again confi rmed similarity indexes. the general irreproducibility associated with biochemical It is well established now that the presence of identifi cation schemes. Consequently, and since the Biolog sphingomonads in drinking water distribution systems system was not useful even to identify those isolates at is not desirable, with some strains being considered as the genus level, we have established a more applicable potential pathogens. Waterborne bacteremia among 16S rRNA gene-based PCR assay that can be easily neutropenic patients in a hospital in Finland was caused employed for the screening of sphingomonads amongst by Sphingomonas paucimobilis (Perola et al. 2002). We colonies growing on R2A agar. The modifi ed primer set have previously reported, in agreement with Koskinen used can be specifi cally employed to differentiate and et al. (2000), that the presence of sphingomonads in confi rm the number of colonies belonging to this genus drinking water may be much more common than has in samples collected from drinking water networks as been reported (Tokajian and Hashwa 2004a; Tokajian et well as storage tanks. Isolates designated as ST-1, 4, 6, al. 2005). The intermittent mode of supply, which is the 9, 10, 13, 16, 17, 25, 29, 44, and 45 were previously strategy employed in all countries suffering from water identifi ed based on 16S rDNA sequencing. ST-4, 6, 9, 10, shortage in the Middle East, is characterized by water and 25 were identifi ed as Sphingomonas natatoria, ST- stagnation and fl ow interruption, that on start up is 13 and ST-45 as Novosphingobium subarcticum, ST-16 associated with biofi lm detachment (Tokajian et al. 2005). as Sphingomonas aquatilis, ST-17 as Sphingomonas sp. The intermittent mode of supply with high and fl uctuating strain IFO 15917, ST-18 as Sphingomonas adhaesiva, ST- pressure causes the frequent detachment of bacterial 29 as Sphingomonas sp. strain MBIC 3990, and ST-44 clumps into the water column, including Gram-negative as Sphingomonas chungbukensis (Tokajian and Hashwa biofi lm–forming bacteria such as Strenotrophomonas, 2004a; Tokajian et al. 2005). Sphingomonas spp., Acidovorax sp., and Pseudomonas It is worth mentioning that ST-4, 9, 10 (Sphingomonas (Tokajian and Hashwa 2004a; Tokajian et al. 2005). natatoria), and isolate ST-18 (Sphingomonas Establishing a rapid and feasible molecular assay for adhaesiva) showed at least two additional bands with the identifi cation of yellow pigmented colonies isolated a size of around 500 bp upon amplifi cation using the from drinking water samples in Lebanon is important sphingomonad-specifi c primer set. This characteristic to minimize health-related risk factors, taking into

254 16S rRNA Gene-Based Identifi cation of Yellow Colonies consideration that phenotypic methods (Biolog system) Condon C, Squires C, Squires CL. 1995. Control of usually fail to generate reproducible phenograms. rRNA transcription in Escherichia coli. Microbiol. Rev. 59:623–645. Conclusion Hsueh R, Teng J, Yang C, Chen C, Pan J, Ho W, Luh T. 1998. Nosocomial infections caused by The abundance and frequent isolation of derivatives of Sphingomonas paucimobilis: Clinical features and yellow pigmented colonies from drinking water samples microbiological characteristics. Clin. Infect. Dis. in Lebanon raised the need to establish a feasible assay to 26:676–681. identify sphingomonads. Amplifi cation of 16S rRNA gene Kabadjova P, Dousset X, Le Cam V, Prevost H. 2002. in sphingomonads, which are known to be likely drinking Differentiation of closely related Carnobacterium water contaminants originating from distribution food isolates based on 16S-23S ribosomal DNA network biofi lms, provides an accurate and sensitive intergenic spacer region polymorphism. Appl. assay for their screening and accurate identifi cation. The Environ. Microbiol. 68:5358–5366. method also overcomes the problems associated with Kalmbach S, Manz W, Wecke J, Szewzyk U. 1999. identifying organisms of environmental origin based on Aquabacterium gen. nov., with description of their metabolic profi les using commercially available Aquabacterium citratiphilum sp. nov., Aquabacterium kits such as the Biolog. Finally, we recommend testing parvum sp. nov. and Aquabacterium commune sp. the feasibility of this assay in identifying sphingomonads nov., three in situ dominant bacterial species from directly from water samples without the need for prior Berlin drinking water system. Int. J. Syst. Bacteriol. cultivation to improve practicability of this assay. 49:769–777. Koskinen R, Ali-Vehmas T, Kampfer P, Laurikkala M, References Tsitko I, Kostyal E, Atroshi F, Salkinoja-Salonen M. 2000. Characterization of Sphingomonas isolates Amy P, Haldeman D, Ringelberg D, Hall D, Russell C. from Finnish and Swedish drinking water distribution 1992. Comparison of identifi cation systems for systems. J. Appl. Bacteriol. 89:687–696. classifi cation of bacteria isolated from water and Lane D, Pace B, Olsen G, Stahl D, Sogin M, Pace N. endolithic habitats within deep subsurfaces. Appl. 1985. Rapid determination of 16S ribosomal RNA Environ. Microbiol. 58:3367–3373. sequence for phylogenetic analyses. 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Relative Body Size Infl uences Breeding Propensity in Fathead Minnows: Implications for Ecotoxicology Testing Procedure

Michael S. Pollock,1 Shelly E. Fisher,1,2 Allison J Squires,1 Robyn J. Pollock,1 Douglas P. Chivers,2 Monique G. Dubé 1,3*

1 Toxicology Centre, University of Saskatchewan, 44 Campus Drive, Saskatoon, SK S7N 5B3 Canada 2 Department of Biology, University of Saskatchewan, Saskatoon, SK S7N 5E2 Canada 3 School of Environment and Sustainability, University of Saskatchewan, Room 230, Law Building, 15 Campus Drive, Saskatoon, SK S7N 5A6

Numerous factors affect the ability or choice of fi shes to breed. For example, studies demonstrate that the appropriate amount of light, temperature, and food must be present before many species will breed. For some species, we are also aware of social factors that affect breeding, such as the size or colour of one’s potential mates. Although studies on mate choice (i.e., choice of one potential mate over another) and factors affecting breeding are extensive, there remain signifi cant gaps in our knowledge with regard to scientifi cally important species. For instance, the fathead minnow (Pimephales promelas), used by numerous researchers as a test subject in reproductive toxicology and behavioural ecology, has well established physical parameters known to facilitate breeding. Conversely, there is very little data describing social factors which may infl uence breeding. The purpose of the current study was to examine some of the factors affecting mate choice in the fathead minnow. Results indicate a consistent relationship between male and female size (length and mass), which can be used to predict the probability of a couple’s breeding potential. Specifi cally, we found that female minnows prefer larger males. In successful pairs there was a greater difference in size between the male and female as compared with unsuccessful pairs. The fi ndings of this study could substantially improve methods for reproductive studies in laboratories or artifi cial streams by decreasing both the number of pairs tested against baseline performance criteria and the time needed to establish actively breeding individuals. This will decrease the cost and increase the effi ciency of future studies, as well as add ecologically interesting knowledge to the literature regarding a scientifi cally important, ubiquitous, and representative North American fi sh species.

Key words: fathead minnow, size assortative reproduction, ecotoxicology, life cycle bioassasy

Introduction needs of the fathead minnow are well known (Ankely et al. 2001), there have been no published studies on Factors that affect the choice or ability of fi shes to breed mate choice or preference for either gender. Given the compose a well studied area of behavioural ecology (Godin importance of the fathead minnow in toxicology, not to and Briggs 1996; Herdman et al. 2004). Some of these mention behavioural ecology in general, this is an area of factors are physical; for example, many species require immediate need. a specifi c light/dark schedule or certain temperature to For many fi sh species, females choose whether or not stimulate breeding (Bhattacharya 1992), while others breeding occurs (Orians 1969). This most likely stems need the appropriate amount or quality of food (Ankley from the difference in energetic cost between male and et al. 2001). Other factors affecting breeding in fi shes female gamete production (i.e., eggs are more expensive can be attributed to mate choice by one or both of the to produce, Dewsbury 1982). In many experimental partners. For instance, male guppies prefer females of a setups, however, randomly allotted breeding pairs larger size (Herdman et al. 2004), while females display are isolated in individual tanks where only one sexual varying preferences for specifi c male colouration patterns, partner is available. Even when both individuals are dependent on the environment (Godin and Briggs 1996). often in an advanced reproductive state, breeding may Although the breadth of knowledge regarding factors not occur. Given this fact, and given that reassortment of affecting the ability or choice of fi sh to breed is extensive, nonbreeding individuals into new pairs often results in there are still major gaps in our knowledge. For example, breeding events, this failure to spawn may be the result the fathead minnow (Pimephales promelas) is a species of mate selectivity. Because males are generally more of fi sh used in numerous behavioural and toxicological indiscriminant than females in the provision of gametes studies examining reproduction (Ankley and Villeneuve (Orians 1969), it is likely that female fathead minnows 2006), yet little is known on minnow mating choice. withhold their eggs when confronted with a male they While the physical parameters of the reproductive judge to be of inferior quality, even in the absence of alternative partners. One factor which may affect female mate choice in * Corresponding author: [email protected] fathead minnows is male body size. A preference for larger

257 Pollock et al. male body size has been documented in female fi shes Materials and Methods over a variety of species (Bisazza et al. 1989; Ptacek and Travis 1997; Kraak et al. 1999; Basolo 2004; Clotfelter Four separate experiments were conducted from 2006 et al. 2006). Large males may be more desirable due to to 2007, each with their own unique experimental their ability to defend resources important for breeding hypotheses related to the examination of chemical and (Breitburg 1987; DeMartini 1988), superior paternal physical stressors on fathead minnow reproduction. The care (Bisazza et al. 1989), greater fertility (Skinner common link across each experiment was in the pre- and Watt 2007), or the ability to protect females from exposure period conducted in the absence of chemical harassment by other suitors (Basolo 2004). Size-related and physical stress. It is these data that are presented mating preferences have not been previously investigated in the current paper; so, although the fi sh were later in the fathead minnow, and due to its common use in exposed to various stressors as part of larger studies, the experiments, such knowledge is important. Factors identical protocols of the pre-exposure phase allowed which determine mate choice and breeding effi ciency in us to use the data to examine the relationship between this species are certainly of interest for fi sheries ecology size and breeding effi ciency. The purpose of the pre- and also have importance for experimental design. exposure period, in relation to the larger studies, was to The purpose of the current study was to examine obtain fi sh which met selection criteria for fertilization size-related mating preference in minnows. A common success, breeding attempts, and survival, which were then protocol used in toxicology (Ankley et al. 2001; Rickwood moved into the exposure phase of the experiment. The et al. 2006a, 2006b, 2006c; Rickwood and Dubé 2007; methods presented below outline the experimental set- Rickwood et al. 2008) calls for a pre-exposure phase up for the pre-exposure phase of each experiment where followed by an exposure phase. The purpose of the pre- breeding success and factors associated with it were exposure phase is to determine baseline reproductive evaluated. A summary of experimental populations used performance in pairs (or groups) and to select which pairs in Experiments 1 to 4 can be found in Table 1. (or groups) meet performance criteria for toxicology testing in the exposure phase of the experiment. This Experiment 1—Laboratory Study means the researcher must start with many more pairs than are actually needed to ensure the required numbers Minnows (12 months old) were purchased from a of breeding couples that meet criteria are obtained. The commercial supplier (Osage Catfi sheries Inc., Osage ambition of the current study is to decrease the number Beach, Missouri). Fish were maintained in the laboratory of pairs that do not breed, thus decreasing the overall in 529-L holding tanks (Living Stream) at 16°C on a 16:8 number of pairs needed, as well as the time necessary to light:dark cycle and fed commercial fi sh fl akes (Tetrafi n obtain successful breeders. Specifi cally, we hypothesize tropical fi sh fl akes) ad libitum. Males and females that the size (length and mass) of male minnows plays were held in the same tank separated by a mesh screen a signifi cant role in whether or not breeding occurs. We (included in purchase of Living Stream, approximately further predict that females prefer larger males (as seen in 5-mm by 5-mm mesh size). The pre-exposure phase of other species) and that the preference will be proportional this study was conducted for 14 days. The experiment to a female’s own size. was initiated by placing 61 measured pairs of male and female fathead minnows randomly into 10-L experimental tanks containing a nesting object (10-cm section of 10-

258 Improved Minnow Reproduction in Ecotoxicology cm diameter pipe cut in half lengthwise) and an airstone. this experiment was conducted as a fi eld exposure over The tanks were held at 25°C with a 16:8 light:dark cycle 21 days on the banks of the Wabigoon River in Dryden, and had a fl ow-through rate of four turnovers per day of Ontario. Minnows were housed in eleven artifi cial stream heated, fi ltered, and dechlorinated tap water. systems (see Rickwood and Dubé 2007 for details). Following recording of weight, length, and presence As in Experiment 1, following recording of weight, of secondary sexual characteristics, males and females length, and secondary sexual characteristics, males and were randomly assigned to tanks and breeding partners, females were randomly assigned to breeding partners, resulting in females being, on average, 88% of the length resulting in females being, on average, 89% of the length of their male partners. The presence and magnitude of of their male partners. Each artifi cial stream system secondary sexual characteristics were scored qualitatively contained eight 10-L circular test aquaria that sat on a by researchers (current study authors) with signifi cant common table and drained into an 80-L mixing reservoir. experience in fathead minnow reproduction and breeding Each mixing reservoir received four turnovers of reference characteristics. The units of measure as reported were river water daily, which was pumped continuously and qualitative and categorical in nature. in a partially recirculating manner into each aquaria. Fish were fed, twice daily, 1.0 g per couple of Aquaria were covered with 500-micron Nitex mesh which brine shrimp (Artemia spp., San Francisco Bay Brand, allowed water to fl ow over the sides while containing Newark, California.) at 0900 h, and 1.0 g per couple the fi sh. Each mixing reservoir contained a heater (set at of bloodworms (Chironomus spp., San Francisco Bay 25°C) and an airstone to ensure suffi cient oxygenation Brand, Newark, California) at 1700 h. Water quality and heat to the aquaria. Each aquarium contained a pair samples were collected from nine random aquaria each of breeding minnows and a spawning tile (10-cm section day (~15% of aquaria) prior to the morning feeding. of 10-cm diameter pipe cut in half lengthwise). Conductivity, hardness, chlorine, pH, temperature, Aside from the differences in experimental set-up, dissolved oxygen, and ammonia levels were recorded fi sh were fed the same amount and type of food each (Table 2). Following the morning feeding, pairs were left day as described for Experiment 1. Eggs were checked for one hour, after which all nesting objects were checked and water quality data (one water quality sample from for eggs. If fertilized eggs were present, the breeding each table per day) were collected in a similar fashion to attempt was considered a success and recorded as such. Experiment 1. Statistical analysis was performed as per Following the experiment, pairs were divided into Experiment 1. two statistical groups: those that bred, and those that did not. All data were analyzed using a two-sample t-test Experiment 3—Laboratory Study conducted with Systat 11.0 (e.g., was there a difference in length between breeding and nonbreeding males Fathead minnows used in this study were obtained from and females?). Prior to testing, all data were tested for a commercial supplier (Thomas Fish Supply, Anderson, assumptions of normality (Kolmorgorov Smirnov [K- California). All fi sh were between 12 and 16 months S] test) and equality of variance (F-test). In the current old. Prior to the experiment, fi sh were held in conditions study, all data sets met normality standards (populations identical to those in Experiment 1. The study was signifi cantly identical to a normal distribution [p > 0.05] conducted for six days and was initiated by placing 61 and had statistically identical variances [p >0.05]). pairs of male and female fathead minnows randomly into 10-L experimental tanks containing a nesting object (10- Experiment 2—Field Study cm by 10-cm polyvinyl chloride pipe cut lengthwise) and an airstone. As in previous experiments, male and female Fathead minnows (16 months old) were obtained from length, weight, and secondary sexual characteristics were the Pulp and Paper Research Institute of Canada (Pointe recorded. Unlike Experiments 1 and 2, pairs of fi sh were Claire, Quebec). All fi sh were held under conditions size-assorted, ensuring that a minimum of 15% difference identical to Experiment 1. The pre-exposure phase of (i.e., males at least 15% larger than females) existed

259 Pollock et al. between male and female length (females 82% of male length). Light cycle, water temperature, feeding schedule and amount, and water quality measurements were conducted as per Experiment 1. An identical protocol to that used in Experiment 1 was used in ascertaining breeding success and egg collection. Statistical analysis was also carried out in a manner identical to Experiments 1 and 2.

Experiment 4—Laboratory Study

Minnows used in the study were purchased from a commercial supplier (Osage Catfi sheries Inc., Osage Beach, Missouri). They were 14 months old and were held in conditions identical to prior experiments both before and during the study. Experiment 4 was conducted in tanks and included materials and protocols identical Figure 1: Mean (+ SE) fork length (mm) of breeding to Experiments 1 and 3. Following the experimental and non-breeding male and female fathead minnows in protocols of Experiment 3, we ensured that a minimum Experiments 1 (breeding couples n = 32, non-breeding 15% difference existed between male and female length couples n = 29) and 2 (breeding couples n = 57, non-breeding (females 84% of male length), and measured and couples n = 32) (* denotes signifi cant differences between statistically tested breeding success. Experiment 4 differed pairs, t-test, D = 0.05). from Experiments 1 to 3 in that there were two females for each male in each aquarium. The purpose of adding the extra female was to determine if breeding rates could be increased when compared with Experiment 3.

Results

Results of Experiments 1 and 2 demonstrated similar trends. Results indicated that longer males were more likely to breed than smaller males (Experiment 1, t-test, p = 0.04; Experiment 2, t-test, p = 0.02, Fig. 1), while no effect of female length on reproduction could be detected (Experiments 1 and 2, t-test, p = 0.11 and 0.23, respectively, Fig. 1). Experiment 2 indicated that heavier males were more likely to breed than lighter males (t-test, p = 0.04, Fig. 2), while Experiment 1 showed no effect of male weight on breeding success (t-test, p = 0.41, Fig. 2). However, in Experiment 1 lighter females were Figure 2: Mean (+ SE) mass (g) of breeding and nonbreeding male and female fathead minnows in Experiments signifi cantly more likely to breed than heavier females 1 (breeding couples n = 32, non-breeding couples n = 29) (t-test, p = 0.005, Fig. 2); again, a result that was not seen and 2 (breeding couples n = 57, non-breeding couples n = in Experiment 2 (t-test, p = 0.28, Fig. 2). 32) (* denotes signifi cant differences between pairs, t-test, The most signifi cant factor in breeding success (i.e., D = 0.05). lowest p-value and largest effect size) was the magnitude of the difference in both length and weight between male and female partners. In both studies, males were those who had not (t-test, Experiment 3, 41 breeders, more likely to breed when paired with relatively smaller 20 nonbreeders, male length, p = 0.45, male mass, p = females (Experiment 1, t-test, p = 0.001; Experiment 2, 0.85; Experiment 4, 42 breeders, 19 nonbreeders, male t-test, p = 0.004, Fig. 3). A similar effect was noted in all length, p = 0.31, male mass, p = 0.79). However, when studies with regard to mass, in that signifi cantly larger female length was compared between breeders and differences were noted in breeding partners compared nonbreeders for Experiment 3, a signifi cant difference with nonbreeding partners (Experiment 1, t-test, p = was noted, with breeding females being signifi cantly 0.005; Experiment 2, t-test, p = 0.007, Fig. 3). shorter than nonbreeding females (fork length = 53.7 In Experiments 3 and 4, which involved direct mm and 55.8, mm respectively; t-test, p = 0.01), with manipulation of the length relationship between males no difference between length in Experiment 4 or mass in and females, we noted no signifi cant difference in either either experiment (t-test, Experiment 3, female mass, p = mass or length of male minnows that had bred versus 0.25; Experiment 4, female length, p = 0.23, female mass,

260 Improved Minnow Reproduction in Ecotoxicology

Figure 4: Rate of breeding (cumulative percentage of couples that bred each day) for Experiments 1 (n = 61), 2 (n = 89), 3 (n = 61) and 4 (n = 61) (lettering indicates similar groups, K-S test, D = 0.05).

Figure 3: Mean (+ SE) ratio of female to male length (%) (female length/male length * 100) and mass (female mass/ four experiments. This was important as longer males male mass * 100) between breeding and nonbreeding are usually older and thus may be more likely to possess pairs of male and female fathead minnows in Experiment advanced secondary sex characteristics compared with 1 (breeding couples n = 32, non-breeding couples n = 29) younger shorter males. So, to ensure it was size (length and Experiment 2 ( breeding couples n = 57, non-breeding and mass) that females were choosing rather than an couples n = 32) (* denotes signifi cant differences between advanced state of breeding appearance, we statistically pairs, t-test, D = 0.05 ). compared the presence of secondary sex characteristics between breeding and nonbreeding males and females. p = 0.36). Similarly, there was no difference in relative Results of the chi-square analysis indicated no signifi cant length or mass between breeding and nonbreeding pairs difference between the secondary sexual characteristics for either Experiment 3 (t-test, length, p = 0.19, mass, p = (presence of fi n dot, mucous pad, tubercles, banding, 0.44) or 4 (t-test, length, p = 0.32, mass, p = 0.18). and ovipositor) of breeding and nonbreeding males or In order to test the hypothesis that size-matching females in any of the four experiments (see Table 3 for all the minnows would increase breeding effi ciency, we comparisons, sample sizes, and p-values). compared the cumulative breeding attempts (total percentage of couples that bred by day X) between all Discussion 4 experiments for the fi rst six days of each study using a K-S test. The total percentage of breeders in Experiment Our results indicate that size, particularly the size ratio 3 was 67%, or a rate of 11.2% (67% per 6 days) per between genders, is a powerful predictor of breeding day, which was similar to the rate in Experiment 4 potential in fathead minnows (Fig. 3). In Experiments 1 (11.5%, or 69% per 6 days) per day. Comparatively, and 2, there was a signifi cant difference in the female Experiment 1 had a rate of 3.79% (52% per 14 days) to male weight to length ratio between breeding and per day, and Experiment 2 had a rate of 3.04% (64% nonbreeding groups. There was also a difference per 21 days). A K-S test revealed a signifi cant difference approaching signifi cance between breeding and between Experiments 1 (randomly assigned pairs) and 3 nonbreeding females in Experiment 1 (p = 0.11, Fig. 1). (15% difference between pairs) (K-S value = 0.83, p = Using this data, Experiments 3 and 4 were conducted 0.005, Fig. 4), as well as Experiments 1 and 4 (trios, 15% and involved deliberate size manipulation, ensuring difference between pairs) (K-S value = 0.67, p = 0.004, that each female was at least 15% smaller in length Fig. 4). Similarly, Experiments 2 (randomly assigned than her male partner (largest consistent and obtainable pairs) differed signifi cantly from Experiments 3 (15% difference in our experimental population). The only difference between pairs) (K-S value = 0.83, p = 0.004, signifi cant difference between breeders and nonbreeders Fig. 4) and 4 (trios, 15% difference between pairs) (K-S in Experiments 3 and 4 was female length. In Experiment value = 0.83, p = 0.004, Fig. 4). However, no difference 3, smaller females were signifi cantly more likely to breed existed when Experiments 3 (pairs, 15% minimum than larger females (Fig. 1); given the fact that all couples between pairs) and 4 (trios, 15% minimum between male were size-matched, mate choice or the choice to breed and both females) were compared (K-S value = 0.17, p = in Experiment 3 may have been affected by traits other 0.50, Fig. 4) or Experiments 1 and 2 (K-S value = 0.5, p than size. = 0.17, Fig. 4). Preferences for large males have been documented To be sure that our results were indicative of male in numerous fi sh species with multiple reasons proposed size, we compared the secondary sex characteristics for this inclination. In some species, defence of resources between breeding and nonbreeding individuals for all necessary for breeding is an important factor in

261 Pollock et al.

reproductive success (Breitburg 1987; DeMartini 1988). over female choice (Burley et al. 1996; Mays and Hill Fathead minnows require a nest site which is guarded 2004; Miller and Brooks 2005; Jawor and Breitwisch by the male (Marcus 1934); therefore, it is possible that 2006) since these characteristics may indicate the larger males are better at securing the most desirable nest presence of quality genes (Mays and Hill 2004). Since sites, as sometimes occurs in other species (Breitburg our females did not have a choice between a male with 1987; DeMartini 1988). Male fathead minnows also well developed secondary sex characteristics and a male perform paternal care by guarding eggs until they hatch with limited ornaments, we cannot directly conclude that (Unger 1983). Larger males may be more successful in they do not play a role in female choice in the fathead defending eggs from intruders, increasing the hatching minnow. However, in our study, it is clear that the more success of eggs (Bisazza et al. 1989) and representing important factor determining mate choice or choice to a direct reproductive benefi t for the female. It has also breed in fathead minnows is size. It would be of interest been demonstrated in some species that larger males are to determine if secondary sexual characteristics also capable of producing greater quantities of sperm (Zbinden impart some importance in mate choice, and what that et al. 2001). If size is in fact an honest indicator of sperm role would be relative to mate size. Studies conducted in production in the fathead minnow, female selectivity which females have the choice between large and small directed towards larger males could result in higher males with and without ornamentation could directly reproductive success. Eggs left unfertilized represent a answer this question. great reproductive waste for a female because eggs are In some species, male choice may also play a role energetically costly to produce (Dewsbury 1982). If small in breeding behaviour. Males of many fi sh species have males are unable to fertilize full clutches of eggs from been shown to prefer larger females, since greater size very large, gravid females, selective pressure would tend may indicate increased fecundity (Herdman et al. to favour females with a preference for large males. 2004). It is improbable however that male mate choice While size appeared to be an important factor in exerted a major role in this experiment. From previous determining whether breeding occurred, secondary studies conducted in our research group (Pollock et al. sexual characteristics had no signifi cant effect (Table unpublished data) as well as others (Cole and Smith 1). This is an interesting occurrence given that male 1987), it is clear that males court females continuously ornamentation is often presumed to have great infl uence regardless of female size, and it is the females that limit

262 Improved Minnow Reproduction in Ecotoxicology the frequency of mating. In the current study, males 2001), if all couples were size-sorted, it would allow for tirelessly pursued any female that approached the nest, a more standardized test method, making comparisons while females most commonly avoided these advances. between studies and laboratories more relevant. Data from Experiments 1 and 2 indicate that Therefore, we encourage researchers involved in the the size ratios which appeared to affect mating were study of reproductive fathead minnows not only to utilize different between experiments (Fig. 3). It is possible that the data presented in the current study (minimum 15% these differences are the result of population variation difference in male to female length ratio), but also to in mate choice. For example, the absolute size of males examine other factors which potentially affect effi ciency and females may differ in the two populations affecting of breeding in laboratory fi shes. Through these efforts the size ratio at which breeding is more likely to occur. we suggest a standardized protocol could be established, This phenomenon has been demonstrated in guppies, in and which could be applied with great success across which preference for male ornamentation differs greatly ecotoxicology. between populations (Brooks 2002). It is also possible that females assessed other males within the laboratory Acknowledgments population prior to the experiment and used this This project was funded by NSERC, Canada Research information when later confronted with a single male. Chairs Program and the Canadian Foundation for Prior to all experiments, holding conditions permitted Innovation (Dr. Monique Dubé). The authors would fi sh to maintain visual and chemical contact through like to thank the editor and reviewers whose time and a mesh divider (included in purchase of Living Stream, comments have added to the quality of our manuscript. approximately 5-mm by 5-mm mesh size), presumably allowing females to assess the overall quality of males References within the population. This may have affected subsequent mate choice. Ankley G, Jensen KM, Kahl MD, Korte JJ, Makynen Along with understanding the dynamics behind EA. 2001. Description and evaluation of short- mating preference in fathead minnows, the understanding term reproduction test with the fathead minnow of conditions that stimulate breeding could potentially (Pimephales promelas). Environ. Toxicol. Chem. save time and money, as well as eliminate important 20:1276–1290. confounding variables from experimental design. The Ankley GT, Villeneuve DL. 2006. The fathead minnow in extensive utilization of the fathead minnow in biological aquatic toxicology: Past, present and future. Aquat. and toxicological studies (i.e., Ankely protocol [Ankely Toxicol. 78:91–102. et al. 2001] and its variants, e.g., Rickwood et al. 2006a, Basolo AL. 2004. Variation between and within the sexes 2006b, 2006c) warrants further research into how in body size preferences. Anim. Behav. 68:75–82. relative body size can infl uence breeding propensity. The Bhattacharya S. 1992. Endocrine control of fi sh results of such research have the potential to be used for reproduction. Curr. Sci. 63:135–139. establishing new guidelines or protocols. Bisazza A, Marconato A, Marin G. 1989. Male If our study, along with similar works, could be competition and female choice in Padogobius martensi used to generate and supplement existing reproductive (Pisces, Gobiidae). Anim. Behav. 38:406–413. protocols (i.e., Ankely protocol), it would offer several Breitburg DL. 1987. Interspecifi c Competition and the advantages to reproductive toxicology. For example, Abundance of Nest Sites: Factors Affecting Sexual many studies use reproductive groups rather than pairs Selection. Ecology 68:1844–1855. to increase the breeding potential. While this approach Brooks R. 2002. Variation in Female Mate Choice within is successful, it has several problems. First, it involves Guppy Populations: Population Divergence, Multiple the use of signifi cantly more test animals. Second, any Ornaments and the Maintenance of Polymorphism. study using multiple females or males in a single tank is Genetica 116:1573–6857. not able to correlate factors such as egg production or Burley NT, Parker PG, Lundy K. 1996. Sexual selection fertilization ability (i.e., impossible to know who laid or and extra pair fertilization in a socially monogamous fertilized the eggs) with other endpoints specifi c to the passerine, the zebra fi nch (Taeniopygia gullata). Behav. Ecol. 7:218–226. fi sh (i.e., hormone level, body condition, etc.). By using Clotfelter ED, Curren LJ, Murphy CE. 2006. Mate Choice well-matched pairs instead of groups, several advantages and Spawning Success in the Fighting Fish Betta could be gained while decreasing the drawbacks of using a splendens: The Importance of Body Size, Display single male and female per tank, which is mainly a decrease Behavior and Nest Size. Ethology 112:1170–1178. in reproductive potential. Advantages would include the Cole KS, Smith RJF. 1987. Male courting behaviour in savings of time, money, and laboratory resources due to the fathead minnow, Pimephales promelas. Environ. the use of fewer test animals in the search for actively Biol. Fishes 18:235–239. reproducing pairs. This would also lead to shorter studies DeMartini EE. 1988. Spawning success of the male due to the rapid and consistent onset of reproduction in plainfi n midshipman. I. Infl uences of male body size well-matched couples. Furthermore, analogous to the and area of spawning site. J. Exp. Mar. Biol. Ecol. use of current standardized protocols (i.e. Ankely et al. 121:177–192.

263 Pollock et al.

Dewsbury DA. 1982. Ejaculate Cost and Male Choice. Skinner AMJ, Watt PJ. 2007. Phenotypic correlates Am. Nat. 119:601–610. of spermatozoon quality in the guppy, Poecilia Godin JGJ, Briggs SE. 1996. Female mate choice under reticulate. Behav. Ecol. 18:47–52. predation risk in the guppy. Anim. Behav. 51:117– Unger LM. 1983. Nest defense by deceit in the fathead 130 minnow, Pimephales promelas. Behav. Ecol. Herdman EJ, Kelly E, Clint D, Godin JGJ. 2004. Male Sociobiol. 13:340–5443. Mate Choice in the Guppy (Poecilia reticulata): Do Zbinden M, Largiader CR, Bakker TCM. 2001. Sperm Males Prefer Larger Females as Mates? Ethology allocation in the three-spined stickleback. J. Fish. 110:97–111. Biol. 59:1287–1297. Jawor JM, Breitwisch R. 2006. Is Mate Provisioning Predicted by Ornamentation? A Test with Northern Cardinals (Cardinalis cardinalis). Ethology 112:888– 895. Received: 18 March 2008; accepted: 25 November 2008. Kraak SB, Bakker M, Theo CM, Mundwiler B. 1999. Sexual selection in sticklebacks in the fi eld: correlates of reproductive, mating, and paternal success. Behav. Ecol. 10:696–706. Marcus HC. 1934. Life History of the Blackhead Minnow (Pimephales promelas). Copeia 1934:116–122. Mays HL Jr, Hill GE. 2004. Choosing mates: Good genes versus genes that are a good fi t. Trends Ecol. Evol. 19:554–559. Miller LK, Brooks R. 2005. The effects of genotype, age, and social environment on male ornamentation, mating behavior and attractiveness. Evolution 59:2414–2425. Orians GH. 1969. On the Evolution of Mating Systems in Birds and Mammals. Am. Nat. 103:589–603. Ptacek MB, Travis J. 1997. Mate Choice in the Sailfi n Molly, Poecilia latipinna. Evolution 51:1217–1231. Rickwood CJ, Dubé MD. 2007. Application of a pair- breeding fathead minnow (Pimephales promelas) adult reproduction bioassay to a pulp mill effl uent. Water Qual. Res. J. Can. 42:82–90. Rickwood CJ, Dubé MG, Weber LP, Lux S, Janz DM. 2007. Assessing effects of a mining and municipal sewage effl uent mixture on fathead minnow (Pimephales promelas) reproduction using a novel, fi eld-based trophic-transfer artifi cial stream. Aquat. Toxicol. 86(2):272–286. Rickwood CJ, Dubé MG, Hewitt ML, Kovacs TG, Parrott JL, MacLatchy DL. 2006a. Use of paired fathead minnow (Pimephales promelas) reproductive test. Part 1: Assessing biological effects of fi nal bleached kraft pulp mill effl uent using a mobile bioassay trailer system. Environ. Toxicol. Chem. 25:1836–1846. Rickwood CJ, Dubé MG, Hewitt LM, Kovacs TG, MacLatchy DL. 2006b. Use of paired fathead minnow (Pimephales promelas) reproductive test. Part 2: Source identifi cation of biological effects at a bleached kraft pulp mill. Environ. Toxicol. Chem. 25:1847–1856. Rickwood CJ, Dubé MG, Weber LP, Driedger KL, Janz DM. 2006c. Assessing effects of metal mining effl uent on fathead minnow (Pimephale promelas) reproduction in a trophic-transfer exposure system. Environ. Sci. Technol. 40:6489–6497.

Distribution of 14C-labelled Atrazine, Methoxychlor, Glyphosate, and Bisphenol-A in Goldfi sh Studied by Whole-Body Autoradiography (WBARG)

Claude Rouleau1,2* and Jagmohan Kohli1

1 National Water Research Institute, 867 Lakeshore Road, P.O. Box 5050, Burlington, Ontario, Canada L7R 4A6 2 Present Address: Maurice-Lamontagne Institute, 850 route de la mer, C.P. 1000 Mont-Joli, Québec, Canada G5H 3Z4

Nonpersistent contaminants represent thousands of chemicals used as pesticides, pharmaceuticals, personal care products, additives, etc. Because of this diversity, the assessment of the environmental risks they may pose for the environment represents a formidable task. Identifi cation of target organs is key information needed to orient further research on newly- investigated organic xenobiotics. We used whole-body autoradiography to visualize the distribution of 14C-labelled atrazine, methoxychlor, glyphosate, and bisphenol-A in goldfi sh (Carassius auratus) and identify target organs. Fish were exposed for 2 days (glyphosate and bisphenol-A) and 7 days (atrazine and methoxychlor) to the radiolabelled compounds at a concentration of 15 nM. They were then frozen, embedded in carboxymethylcellulose gel, 20-Pm-thick cryosections were collected, freeze-dried, and exposed to phosphor screens to visualize the tissue distribution of radioactivity. Goldfi sh did not accumulate glyphosate. The three other compounds were accumulated, mostly in the gall bladder. Nevertheless, unforeseen accumulation sites were observed; atrazine accumulated in the uveal tract of the eye, high levels of radioactivity were found in the cerebrospinal fl uid of goldfi sh exposed to methoxychlor, and an important accumulation of bisphenol-A was seen in urine, oral mucosa, esophagus, and intestinal lumen. The potential toxicological consequences of the accumulation of these chemicals at very specifi c locations within the fi sh body are discussed and further research suggested.

Key words: atrazine, bisphenol-A, methoxychlor, tissue distribution, fi sh, whole-body autoradiography

Introduction complicated by the fact that very low levels of certain xenobiotic compounds, such as endocrine disrupters, There is increasing evidence that trace amounts of may perturb the normal biochemical balance of fi sh, many household and industrial chemicals, such as birds, and mammals (Crews et al. 1995), requiring the organophosphates, chlorinated pesticides, phenolic use of costly and time-consuming analytical chemistry. compounds, and synthetic estrogens, can perturb the Whole-body autoradiography (WBARG) (Ullberg endocrine system of aquatic organisms and may lead 1954; Ullberg et al. 1982) allows rapid and precise to reproductive failure (Islam and Tanaka 2004). This visualization of the distribution of a radiolabelled stresses the need to thoroughly and adequately assess the chemical in all the organs and tissues of a whole animal. environmental risks they may pose. In view of the very Though it is commonly used in the pharmaceutical large number of chemicals in use today and the fact that industry to study the distribution of new drugs (Solon hundreds of new ones are marketed every year, the task et al. 2002), its use in environmental studies is not of comprehensively assessing the links between chemical widespread. Nevertheless, WBARG has proven its value releases, environmental concentrations, target organism by enabling researchers to illuminate unsuspected routes exposures, tissue concentrations, and probability of of accumulation of metals, organometals, and organic adverse effects represents a formidable challenge to the chemicals in fi sh (Solbakken et al. 1984; Tjälve et al. scientifi c community (MacLeod et al. 2004). 1986, 1988; Bernhoft et al. 1994; Rouleau et al. 1999, One of the key parameters needed by scientists who 2003; Ruus et al. 2001; Scott et al. 2003). want to assess the environmental risks of a given chemical Here we present the results of a preliminary is the identifi cation of the preferential accumulation experiment aimed at identifying target organs and tissues site(s) in aquatic biota following exposure since it is the that accumulate chemicals (and/or their metabolites) in combination of a chemical’s selective accumulation (Tsai goldfi sh (Carassius auratus) upon exposure via water. and Liao 2006) at a specifi c target site and the mode of Goldfi sh has been extensively used as a biological model action at that site that determine the likelihood of toxic in neurophysiology (Finger 2008), and also to assess the effects (MacLeod et al. 2004). Identifi cation is often toxicity of chemicals in the aquatic environment (Chen et al. 2005; Teather and Parrot 2006; Liu et al. 2007; Yin et al. 2007). Chemicals examined were atrazine, a * Corresponding author: [email protected] triazine broadleaf herbicide (Solomon et al. 1996; Allran

265 Rouleau and Kohli and Karasov 2000, 2001), methoxychlor, a chlorinated Biochemicals. Radiochemical purity was at least 95% pesticide used as a substitute to DDT (Johnson and Finley and the radiolabelled chemicals were used without 1980; Magliulo et al. 2002; ATSDR 2004; Versonnen et al. further purifi cation. Goldfi sh (10- to 15-g body 2004), glyphosate, perhaps the most important herbicide weight) were bought from Aleongs International Inc. ever developed (EXTOXNET 1996; Baylis 2000), and (Mississauga, Ontario). They were kept in 20-L aquaria bisphenol-A, used in the production of epoxy resins fi lled with city of Burlington (Ontario) tap water that and polycarbonate plastics and a potent estrogen mimic was previously dechlorinated with activated charcoal (Routledge and Sumpter 1996; Hing-Biu and Peart 2000; and continuously aerated. The aquaria were maintained Belfroid et al. 2002; Pait and Nelson 2003; Stoker et al. at 20 ± 1oC and under a 12 h-light:12 h-dark photoperiod 2003; vom Saal and Hughes 2005). Molecular structure, in a temperature-controlled room at the National Water water solubility, KOW, and some toxicological data are Research Institute in Burlington. Fish were acclimated shown in Fig. 1 and Table 1. to experimental conditions for one week before the beginning of the experiment, with the water changed Material and Methods three times a week. Fish were not fed during this period nor during the experiment. Atrazine-[ring-14C(U)] was purchased from Sigma- Water in exposure aquaria was spiked with the Aldrich, methoxychlor-[ring-14C(U)], and glyphosate- radiolabelled compounds to a concentration of 15 nM. [glycine-2-14C] from American Radiolabeled Chemicals, The resulting 14C levels varied with the specifi c activity and bisphenol-A-[propyl-2-14C] from Moravek (Table 1) of the labelled compounds. A group of four fi sh

Fig. 1. Molecular structure of atrazine, methoxychlor, glyphosate, and bisphenol-A. * shows the location of the 14C-label.

266 Tissue Distribution of Chemicals in Fish by WBARG

Fig. 2. Radioactivity levels in exposure water, expressed as percentage relative to the radioactivity measured at time 0, as a function of time. was introduced into each aquarium 30 min after spiking. of hexane and dry ice. They were then quickly embedded Since the specifi c activity of radiolabelled atrazine in a carboxymethylcellulose gel on a microtome stage, and methoxychlor was lower than for the two other and the assembly was frozen in the same way. Fifteen to compounds, fi sh were exposed to these two chemicals 20 pairs of 20-Pm-thick cryosections were taken at -20oC for 7 days. Fish in aquaria containing glyphosate at different levels in the body of each fi sh with a specially and bisphenol-A were exposed for 2 days. During the designed cryomicrotome (Leica CM3600). The sections exposure period, aquaria were protected from light to were then freeze-dried and applied to fl exible storage prevent hydrolysis (especially for glyphosate), water was phosphor screens (Perkin-Elmer) for one to two weeks. not changed (temperature = 20oC), and radioactivity in After exposure, the phosphor screens were scanned 2-mL water samples was quantifi ed by liquid scintillation with a Cyclone Storage Phosphor Imager (Perkin- spectrometry using a Packard Tri-Carb Liquid Scintillation Elmer). Autoradiograms obtained were visualized, and Counter Model 2300 TR (see Fig. 2 for sampling radioactivity in tissues of interest was semiquantifi ed frequency). At the end of the exposure period, fi sh were using the software Optiquant (Packard Biosciences) (Fig. submitted to lethal anaesthesia (MS-222, 100 mg/L), 3). briefl y rinsed in clean water, and fl ash-frozen in a slurry

Fig. 3. Autoradiogram showing some of the regions of interest selected for quantifi cation. Numbers between brackets in Fig. 4 to 7 are radioactivity concentrations expressed as digital light units (DLU) per mm2 of section surface. Detection limit was set as 3 times the standard deviation of the average background measured in 10 blank areas on the same phosphor screen. Detection of radioactivity with phosphor screens had a dynamic range extending over 5 orders of magnitude, whereas shades of grey that can be visually distinguished range over some 2 orders of magnitude. Thus, black areas seen in some autoradiograms are not saturated but simply results from the output range used for prints, which was chosen to show as much details as possible. See Table 2 for abbreviations.

267 Rouleau and Kohli

necessity to have enough radioactivity in fi sh to be able to visualize distribution by autoradiography. Fish were not fed to keep ammonia excretion low and to avoid ingestion of food particles or faeces to which radiolabelled chemicals may have adsorbed. Autoradiograms presented (Figs. 4 to 7) show the distribution of both the radioactive parent compound and its metabolites. Features described are typical of those seen in all fi sh of a given exposure group. None of the fi sh died during the experiment and they did not exhibit any behavioural signs of stress. Figure 2 shows the level of 14C radioactivity in water, expressed as a percentage of the activity measured at time 0. Radioactivity level in the water containing atrazine and glyphosate showed little variation (<5%), whereas it decreased by 20% in the water containing bisphenol-A. These tendencies likely refl ect the different uptake of the chemicals by fi sh; no radioactivity could be detected in the fi sh exposed to glyphosate, labelling of atrazine- exposed fi sh was very low, except for the gall bladder, and fi sh exposed to bisphenol-A exhibited a much stronger labelling (see below). In the case of methoxychlor, 14C activity in water decreased by almost 70% during the fi rst 24 h of exposure and increased thereafter to a level representing 65% of the activity measure at time 0. It has been shown that methoxychlor is effi ciently transformed to its monodemethylated and bisdemethylated metabolites by fi sh liver microsomes (Schlenk et al. 1997, 1998). The variation in the 14C activity in water to which Results and Discussion methoxychlor was added may have been caused by the fast uptake of methoxychlor followed by the release of Exposure conditions were a compromise between realistic water-soluble metabolites, conjugated or not. environmental conditions (e.g., concentration in water), Glyphosate did not bioaccumulate in goldfi sh, as minimization of fi sh stress (length of exposure), and the the autoradiograms were completely blank (not shown).

B A

Fig. 4. Autoradiograms and corresponding tissue sections of goldfi sh exposed to 14C-atrazine. Black areas contain the highest concentration of radioactivity. Bar = 1 cm. Autoradiogram and tissue section pairs in A and B are from different fi sh. See Table 2 for abbreviations.

268 Tissue Distribution of Chemicals in Fish by WBARG

our experiment and compared with the concentrations observed in well (Smith et al. 1996) and drainage waters (Vereecken 2005) following land application. The three other chemicals were accumulated to various extents. WBARGs and the corresponding tissue sections are presented in Fig. 4 to 7. In fi sh exposed to atrazine for 7 days, the radioactivity was mainly concentrated in the gall bladder (Fig. 4), indicating that atrazine was readily metabolized. A diffuse and weak labelling can be seen in all the other tissues. Labelling of organs such as the liver, intestine, kidney, and eye lens was somewhat higher. This is similar to the distribution of atrazine in the whitefi sh (Coregonus fera [Gunkel and Streit 1980]), carp (Cyprinus carpio [Gluth et al. 1985]), tilapia (Tilapia sparrmanii [du Preez and van Vuren 1992]), and larvae of the amphibian Xenopus laevis (Edginton and Rouleau 2005). An interesting observation is the labelling of the uveal tract of the eye (Fig. 5). This layer of the eye contains melanin, a polyanionic pigment which is well known to bind numerous substances via ionic interactions, especially positively charged organic amines and metals (Larsson 1993, and references therein; Leblanc et al. 1998; Bridelli et al. 2006). The potential impact of the binding and accumulation of xenobiotics in the eye is still subject to debate. There are indications that xenobiotic binding to melanin may be the main factor in the etiology of chronic lesions affecting melanin-containing tissues (Larsson 1993; Jaga and Dharmant 2006), though Leblanc et al. (1998) argued that melanin binding and retinal toxicity are separate phenomena that are not necessarily related. Nevertheless, the accumulation of atrazine in the uveal tract of goldfi sh raises the question of possible toxic effects on the retina due to the close proximity of these ocular tissues, a question that is certainly worthy of further investigation by environmental toxicologists. In the case of methoxychlor, our data (Fig. 6) show that the gall bladder and intestine contain most of the radioactivity, indicating an effi cient metabolisation and excretion. Most other tissues showed a weak labelling. However, some muscle tissue types show a higher level of radioactivity. The red muscle layer under the skin and the bundle of red muscle fi bres forming the ventral infracarilanis (VI) contained more radioactivity than the white muscle (Fig. 6B). Labelling of the muscles belonging Fig. 5. Detail from autoradiogram and tissue section shown to the adductor mandibulae (AM) complex was also in Fig. 4B. Bar = 0.25 cm. See Table 2 for abbreviations. higher (Fig. 6A, 6B); this may be related to their higher vascularization and different metabolism (Ostrander 2000) compared with white muscle. But the most This may be due to the short exposure time. However, striking feature of methoxychlor disposition in goldfi sh in view of the high water solubility and ionic character was the high labelling of the cerebrospinal fl uid around of glyphosate at environmentally relevant pH (pKa the brain, which was only surpassed by that of the gall = < 2, 2.6, and 5.6) (Mamy et al. 2005), it was not bladder, whereas the labelling of the brain itself was very expected to bioaccumulate or represent a toxic threat low. This indicated that methoxychlor or a metabolite (WHO 1994 and references therein). Adverse biochemical can cross the blood-cerebrospinal fl uid barrier (BCSFB) and histopathological effects of glyphosate have been but not the blood-brain barrier (BBB). Previous work has observed in fi sh (Szarek et al. 2000; Jiraungkoorskul et shown that some metals (Hg, Cd, Mn, Zn) can reach the al. 2003; Glusczak et al. 2006), but at concentrations brain via axonal transport after uptake in water-exposed in water that were 102 to 105 higher compared with sensory organs, thus circumventing the BBB (Rouleau et

269 Rouleau and Kohli

Fig. 6. Autoradiograms and corresponding tissue sections of goldfi sh exposed to 14C-methoxychlor. Bar = 1 cm. Autoradio- gram and tissue section pairs in A and B are from different fi sh. See Table 2 for abbreviations.

Fig. 7. Autoradiograms and corresponding tissue sections of goldfi sh exposed to 14C-bisphenol-A. Dotted box in A shows an enlargement of a section of the autoradiogram with output levels adjusted to reveal labelling differences in gall bladder, intestine, and urine. Bar = 1 cm. Autoradiogram and tissue section pairs in A and B are from different fi sh. See Table 2 for abbreviations.

270 Tissue Distribution of Chemicals in Fish by WBARG al. 1995, 1999; Persson et al. 2003; Scott et al. 2003). bisphenol-A. The phenomenon underlying the higher Tributyltin (TBT) has been shown to cross the BBB, but labelling of the oral mucosa and esophagus, as compared not the BCSFB, and is also taken up in the brain via with skin, remain unknown at present. Nevertheless, axonal transport (Rouleau et al. 2003). It is the fi rst time binding sites in these tissues appear to have quite a high that we observed a preferential uptake of a xenobiotic in affi nity for this compound. Further research is needed to the cerebrospinal fl uid (CSF) of a fi sh. characterize these binding sites and determine whether or The maintenance of a strict homeostasis in the not this may have an impact on fi sh health. central nervous system is a necessary requirement for the highly specialized brain cells to fulfi ll their physiological Conclusion functions. The role of the BBB and the BCSFB in the maintenance of this homeostasis has been extensively WBARG has allowed the identifi cation of unforeseen studied (Suzuki et al. 1997; Zheng et al. 2003; Strazielle accumulation sites in goldfi sh for atrazine (uveal tract), et al. 2004; Löscher and Potschka 2005). The CSF fulfi lls methoxychlor (CSF), and bisphenol-A (oral mucosa). many mechanical, transport, and buffering functions as These fi ndings will help us to orient and focus further well as neuroimmune regulation and transmission of research work on the fate and effects of these chemicals neuroactive compounds. Choroid plexuses, located in in fi sh on the accumulation sites observed. Work will brain ventricles, form the interface between blood and be continued to quantify the biokinetics (rates of CSF. The BCSFB is constituted by the tight junctions uptake, distribution, metabolism, elimination) of these between the epithelial cells that restrict the paracellular compounds in relation to their accumulation in target route, and thus the entry of polar compounds that are not tissues and determine if this can lead to physiological a substrate for transbarrier transporters (Strazielle et al. function impairment under realistic environmental 2004). However, lipophilic compounds of low to medium conditions. molecular weight can cross epithelial cell membranes via passive diffusion (Zheng et al. 2003). Acknowledgment Though methoxychlor has been shown to inhibit brain mitochondrial respiration and increase hydrogen The authors gratefully acknowledge the skillful technical peroxide production in rat and mice, both in vitro and assistance of G. Pacepavicius. in vivo (Schuh et al. 2005), the risk represented for fi sh brain functions is unknown. This risk is also dependent References upon whether it is methoxychlor itself that crosses the BCSFB and accumulates in the CSF or one of its Allran JW, Karasov WH. 2000. Effects of atrazine and metabolites. Methoxychlor is metabolized in the liver nitrate on northern leopard frog (Rana pipiens) larvae exposed in the laboratory from post hatch by O-demethylation to polar mono- and bisphenolic through metamorphosis. Environ. Toxicol. Chem. metabolites (Bulger et al. 1978). The bisphenolic 19:2850–2855. compound, 2,2-bis-(p-hydroxyphenyl)-1,1,1-trichloro Allran JW, Karasov WH. 2001. Effects of atrazine on ethane (HPTE), is known to be a much more potent embryos, larvae, and adults of anuran amphibians. endocrine disrupter than methoxychlor itself (Gaido Environ. Toxicol. Chem. 20:769–775. et al. 1999). Further research is needed to characterize ATSDR. 2004. ToxFAQs for Methoxychlor. Available the biodynamics, the speciation, and the mechanism by on-line at: http://www.atsdr.cdc.gov/tfacts47.html. which methoxychlor crosses the BCSFB. [Accessed: February 2008]. The distribution of bisphenol-A in goldfi sh is Baylis AD. 2000. Why glyphosate is a global herbicide: characterized by a very high labelling, in decreasing strengths, weaknesses and prospects. Pest Manag. order, of urine > gall bladder > intestinal content >> liver Sci. 56:299–308. (Fig. 7). The higher labelling of urine compared with the Belfroid A, van Velzen M, van der Horst B, Vethaak D. gall bladder (Fig. 7A) suggests that urinary excretion is 2002. Occurrence of bisphenol A in surface water more important than the hepatobiliary pathway. In rats, and uptake in fi sh: evaluation of fi eld measurements. bisphenol-A elimination occurred mostly via urinary Chemosphere 49:97–103. excretion of the glucuronic acid conjugate whereas Bernhoft A, Hektoen H, Skaare JU, Ingebrigtsen excretion of the parent compound proceeded via the K. 1994. Tissue distribution and effects on hepatobiliar pathway (Snyder ,et al. 2000; Kurebayashi hepatic xenobiotic metabolising enzyme of et al. 2005). Both bisphenol-A and its glucuronated 2,3’,3’,4,4’-pentachlorobiphenyl (PCB-105) in cod conjugate have been found in fi sh (Lindholst et al. 2001, (Gadus morhua) and rainbow trout (Oncorhynchus 2003). Despite the high radioactivity level in urine, the mykiss). Environ. Pollut. 85:351–359. labelling of the kidney is rather low indicating a fast Bond CE. 1979. Biology of fi shes. Saunders College turnover. The oral mucosa, esophagus (Fig. 7B), and, Publishing, Philadelphia. to a lesser extent, skin are also highly labelled. 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Received: 23 January 2008; accepted: 22 July 2008.

Exposure to Model-Scale Sewage Treatment Plant Effl uent Affects Circulating Sex Steroids in Rainbow Trout

Joanne L. Parrott,1* Mark E. McMaster ,1 Subhash Verma,2 and David Trowbridge2

1Water Science and Technology Directorate, National Water Research Institute, Environment Canada, 867 Lakeshore Rd., Burlington, Ontario, Canada L7R 4A6 2Sault College, Sault Ste. Marie, Ontario, Canada P6A 5L3

Reproductive steroids were assessed in immature rainbow trout (Oncorhynchus mykiss) after 21-day exposures to 10% primary or 100% secondary-treated model-scale sewage treatment plant (STP) effl uent. Plasma testosterone was elevated over 4-fold in rainbow trout exposed to 10% primary model STP effl uent and 100% secondary model STP effl uent, and 1.7- fold and 2.5-fold in trout exposed in the following years, 2 and 3, respectively. Exposure to the positive control compound, 17β-estradiol (100 ng/L), raised plasma estradiol concentrations of exposed trout, but had few effects on plasma testosterone concentrations. There was no induction of ethoxyresorufi n-O-deethylase (EROD) activity by any treatment, but liver-somatic indices were elevated in year-1 fi sh exposed to 10% primary or 100% secondary-treated model STP effl uent. The results show that exposure to this model STP effl uent can increase circulating testosterone concentrations. The rainbow trout 21-day test proved to be a simple bioassay that holds promise for onsite assessments of effl uents.

Key words: municipal wastewater effl uent, sex steroids, testosterone, estradiol, mixed function oxygenase (MFO), rainbow trout

Introduction four river estuaries in the U.K. had increased plasma vitellogenin (Vtg), a yolk-precursor protein found only Estrogenic Responses in Wild Fish Exposed to in mature female fi sh (Allen et al. 1999). Feminization Sewage Efﬂ uents of some fi sh was apparent, with 17% of male fl ounder from the Mersey River having developed oocytes in their During the past decade there have been many reports of testes (Allen et al. 1999). Input to the rivers was largely estrogenic responses in fi sh exposed to sewage treatment municipal wastewater effl uents (MWWEs), many of plant (STP) effl uents (Purdom et al. 1994; Jobling et which received only primary treatment at the time of the al. 1998). Wild male roach (Rutilus rutilus) captured fi eld survey in 1996. downstream of several sewage treatment plants in the U.K. had increased vitellogenin, a yolk-protein precursor Estrogenic Responses of Lab Fish and Caged Fish (normally found only in female fi sh) that is produced in Exposed to Sewage Efﬂ uents response to estrogens (Jobling et al. 1998). Histological examination of the testes showed oocytes within testicular Controlled short-term laboratory exposures of fi sh to tissue, a condition termed “intersex” (Jobling et al. STP effl uents have shown estrogenic effects. Juvenile 1998). The incidence of intersex was as great as 100% in rainbow trout (Oncorhynchus mykiss) caged for 2 weeks males from two rivers, despite sampling locations being downstream of a Swedish STP had massive induction of several kilometres from STP effl uent outfalls (Jobling et plasma vitellogenin (1.5 mg of Vtg/mL of plasma) (Larsson al. 1998). Although genetic sex of the roach could not et al. 1999). Rainbow trout held in 100% STP effl uent be determined, it was thought that the intersex fi sh were from Uppsala, Sweden, in 1998 had plasma vitellogenin genotypic males that had been feminized by exposure to concentrations that were four orders of magnitude higher environmental estrogens, based on two lines of evidence: than control fi sh (Deutsch 2000, in Norman et al. 2000). There was an inverse relationship between the proportion After upgrading the nitrifying capacity and doubling the of normal males and intersex males at each site, and retention time of wastewater within the Uppsala STP in there was a signifi cant correlation between the plasma 1999, no induction of Vtg was seen in trout and zebrafi sh vitellogenin and the intersex index of each male (Jobling exposed to 100% STP effl uent (Norman et al. 2000). et al. 1998). Research has shown that the concentrations of Other studies in the U.K. have found similar estrogens in sewage effl uent are suffi cient to cause the feminization of male fi sh captured downstream of STPs. observed Vtg-induction response in fi sh. Effl uents from Male fl ounder (Platichthys fl esus) from nine sites along seven U.K. STPs (treating domestic sewage) were found to be estrogenic (in the yeast estrogen screen, or YES assay, containing a human estrogen receptor), and a * Corresponding author: [email protected] majority of the potency was due to three components:

275 Parrott et al.

17β-estradiol (E2), estrone (E1), and 17α-ethinylestradiol are usually static renewal tests lasting from a week to a (Desbrow et al. 1998). Dramatic increases in plasma Vtg month. The steroids measured in blood plasma or serum of male rainbow trout and roach were seen after 21 included testosterone (T), 11-ketotestosterone, and E2. days exposure to concentrations of E2 and E1 that were Assessment of reproductive steroids in fi sh exposed found in typical STPs in the Desbrow et al. (1998) study to pulp mill effl uents has been shown to be a useful (Routledge et al. 1998). and predictive short-term bioassay. Lower levels of circulating steroids were seen in various species of fi sh Estrogenic Responses in Fish Exposed to North after short-term exposure to Canadian pulp mill effl uents American Sewage Efﬂ uents (McMaster et al. 1996a, 1996b; Parrott et al. 1999a, 1999b; Tremblay and Van Der Kraak 1999; Dubé and The effects of STP effl uents were predicted to be less MacLatchy 2000, 2001a, 2001b). Short-term exposures severe in North America, due to increased fl ow of most of an estuarine fi sh, mummichog (Fundulus heteroclitus) North American rivers compared with rivers in the U.K. were able to detect endocrine-disrupting substances in However, most studies of fi sh exposed to North American pulp mill process streams (Dubé and MacLatchy 2001a). MWWEs have reported similar estrogenic responses to Short-term exposures of rainbow trout to wood-related those seen in fi sh exposed to European STP effl uents. compounds, such as β-sitosterol, have shown that these Male carp (Cyprinus carpio) captured in the Minnesota plant sterols can affect fi sh steroids (Tremblay and Van River downstream from the St. Paul Metropolitan Der Kraak 1999). sewage treatment plant had increased serum Vtg (with The objectives of the present study were to expose no increase in serum E2 concentrations) and decreased rainbow trout for 21 days to primary-treated and serum testosterone concentrations (Folmar et al. 1996). secondary-treated model STP effl uent, and to assess A later study showed similar changes in a second species, impacts on reproductive steroid concentrations. We walleye (Stizostedion vitreum), captured near the St. Paul wanted to assess the potential reproductive impact of a Metropolitan sewage treatment plant (Folmar et al. 2001). Canadian primary and secondary-treated model-scale STP Male and female walleye had increased serum E2 and effl uent. Primary-treated sewage is usually much higher in Vtg, and male walleye had decreased serum testosterone ammonia, total suspended solids, and biological oxygen (Folmar et al. 2001). Two Mississippi MWWEs induced demand parameters compared with secondary-treated Vtg in male channel catfi sh (Ictalurus punctatus) caged effl uent. Because of this we could expose fi sh to only for three weeks in receiving-stream water (Tilton et al. 10% primary-treated model STP effl uent, and to 100% 2002). Studies of male fathead minnows (Pimephales secondary-treated model STP effl uent. Concentrations of promelas) exposed to a Denton (Texas) MWWE as it circulating steroids (T and E2) and hepatic mixed function passed through constructed wetlands have shown Vtg oxygenase (MFO) activity (measured as ethoxyresorufi n- induction and decreases in secondary sex characteristics O-deethylase [EROD]) were measured in immature after 3 weeks exposure (Hemming et al. 2001). Methanol rainbow trout exposed to sewage effl uent (10% primary- extracts of MWWE from Denton (Texas) and Red Hook treated effl uent or 100% secondary-treated effl uent), (New York) induced plasma Vtg of male Japanese medaka E2 (positive control compound, 100 ng/L), or control (Oryzias latipes) after only 7-day exposures (Huggett et water. al. 2003). Male longear sunfi sh (Leopomis megalotis) captured downstream of an Oklahoma MWWE had Materials and Methods increased Vtg and increased plasma testosterone compared with fi sh from a reference site (Porter and Janz Model-Scale Sewage Treatment Facility 2003). Adult male mummichogs (Fundulus heteroclitus) had increased plasma and hepatic Vtg after a 21- The model-scale STP facilities at Sault College in Sault day exposure to 75% Yonkers (New York) MWWE Ste. Marie, Ontario, Canada, were designed to provide (McArdle et al. 2000). Larval sunshine bass (Morone training for those who intended to work in the fi elds of saxatilis x Morone chrysops) exposed to three New York water and wastewater treatment. The fi ve litre per minute City MWWEs for only 4 days had increased expression model-scale STP was acquired from the Wastewater of estrogen receptors and increased Vtg (Todorov et al. Treatment Centre of the Canada Centre for Inland Waters 2002). (CCIW) in Burlington, Ontario. The plant is capable of demonstrating primary, secondary, and tertiary treatment Measurement of Circulating Steroids in Fish Exposed for both municipal and industrial waste waters. to Sewage Efﬂ uents Wastewater was drawn directly from the municipal sewer main (of the town of Sault Ste. Marie, Ont.) by means Although most studies of fi sh exposed to MWWEs have of a submersible grinder pump. Industrial wastewater for focused on measuring Vtg, the measurement of circulating treatment was stored in two 10,000-L insulated tanks. sex steroids in fi sh is another useful way of assessing the Raw surface water was stored in a 20,000-L insulated potential reproductive impacts of MWWEs. Laboratory tank located in the outside compound adjacent to the exposure studies for assessment of steroid levels in fi sh plants.

276 Model STP Effl uent Affects Sex Steroids in Trout

During the exposure the plant was confi gured for Fish were killed by cervical dislocation, weighed, conventional treatment. Primary clarifi cation (340-L and measured (fork length). Fish were dissected and tank) was followed by aeration in a single aeration tank their gonads were visually assessed, and fi sh were sexed (2,800 L) and a fi nal clarifi cation tank (530 L). Typical as female, immature, or male. Most fi sh were female or removal effi ciencies for biological oxygen demand (BOD) immature: Year-1 had 2 male fi sh of 54 and year-3 had 15 and total suspended solids (TSS) are 35% for BOD and male fi sh of 56 total fi sh. The male fi sh were removed from 60% for TSS in the primary and 85% for both BOD and the statistical analyses because plasma T concentrations TSS in the secondary treatment. We do not have the data were very high in these early-developing males. Livers for the time of the exposure, but many years of plant were removed, weighed, and frozen on liquid nitrogen operation have produced these results. (and later stored at -80°C) for later determination of MFO activity, measured as EROD activity. Fish Exposures Livers were removed from the -80°C freezer and thawed on ice. Liver samples were individually Rainbow trout were obtained a from local hatchery homogenized in a HEPES (N-2-hydroxyethylpiperazine- (Iron Bridge, Sault Ste. Marie, Ontario). One replicate N’-2-ethanesulfonic acid) grinding buffer. Liver was performed (n = 10 fi sh per treatment) each year for homogenates were centrifuged at 9,000xg for 20 three years. Year-1 fi sh measured 13.8 ± 1.1 cm (fork minutes and the supernatant (S9) was removed for length, mean ± standard deviation) and weighed 28.2 ± EROD assay. Activity of EROD was measured using a 8.8 g. Years-2 and -3 fi sh were larger: year-2, 20.7 + 1.3 96-well plate kinetic assay that followed the reduction cm, 92.8 ± 14 g; year-3, 20.3 ± 0.98 cm, 86.3 ± 14 g. of 7-ethoxyresorufi n to resorufi n over 12 minutes, using Exposures were done on site at Sault College, close to the a multi-well-plate-reading fl uorometer (Cytofl uor 2300, model-scale sewage treatment facility. Fish were held, 10 Millipore Ltd.; 530-nm excitation fi lter; 590-nm emission per pail, in clean garbage pails with lids (70-L size, fi lled fi lter; sensitivity, 3) (Hodson et al. 1996). Protein content to 50 L) in aerated, charcoal fi ltered, Sault Ste. Marie of the S9 supernatant was determined by the BIO-RAD city water. Temperature was maintained at 12 to 13°C spectrophotometric method (BIO-RAD, Hercules, Calif.), by holding pails in a large circulating water bath. The and specifi c EROD activity was expressed as picomoles of photoperiod was 16 h light to 8 h dark. Fish were fed resorufi n produced per milligram of protein per minute 2% of their body weight per day of standard trout chow (pmol/mg/min). (fi sh food was obtained from the hatchery to ensure Rainbow trout blood samples were taken to maintenance on the same diet for the 21-day exposures). determine levels of circulating sex steroids. T and Fish were fed four hours prior to solution changeover. E2 were determined in blood plasma (pg/ml) using a Exposures were to 10% primary effl uent, 100% radioimmunoassay technique (McMaster et al. 1992, secondary effl uent, high and low E2 (E2 high = 100 ng/L 1995). Blood plasma samples were thawed on ice and [1 mL of stock 5.00 mg E2/mL ethanol in a 50-L pail] analyzed for concentrations of E2 and T. Briefl y, samples and E2 low = 25 ng/L [0.25 mL of stock 5.00 mg E2/mL were extracted with solvent and charcoal, and precipitated ethanol in a 50-L pail]), a water control, and an ethanol in a solvent bath with dry ice. Extracts were tested for solvent control (1 mL ethanol in a 50-L pail, 0.002% ability to compete with radiolabelled E2 or T (Amersham ethanol in fi sh exposure water). Ethanol solvent control Life Science Inc., Arlington Heights, Ill., U.S.A.) for sites fi sh were added to the experimental design, because on an E2- or T-antibody (ICN Biomedicals Inc., Aurora, ethanol was the carrier solvent for the E2 exposures. Fish Ohio, U.S.A.). T and E2 binding curves were run at the exposures were static-renewal, with solutions changed same time as the sample. Results were expressed as pg of three to four times per week (every 2 days). Water quality E2 or T per mL of blood plasma. parameters were measured daily: temperature (10°C), pH (7 to 8.5), dissolved oxygen (>80% saturation), and Data Analyses and Statistics ammonia (0.6 to 1.3 mg/L). Plasma T, E2, and liver EROD activities were analyzed Fish Sampling using analysis of variance (ANOVA) followed by protected two-sample t-tests (SYSTAT, Evanston, Ill.) to After 21 days, fi sh were anaesthetized in a solution of assess Bonferroni adjusted probabilities of differences tricaine methane sulfonate (MS 222, 1 mg/L, Sigma, St. (separate variances) between model STP effl uent-exposed Louis, Mo.) and a blood sample was taken from the dorsal fi sh and control fi sh (exposed to charcoal fi ltered Sault vein through the caudal peduncle (using a heparinized Ste. Marie city water). Differences among fi sh exposed syringe, 1 to 3 mL, 22 to 23 gauge). Blood was iced and to E2 (25 ng/L and 100 ng/L) were compared with the spun in a chilled centrifuge (4°C). Plasma was transferred solvent control fi sh (fi sh exposed to 0.002% ethanol), to pre-iced cryovials (each sample was split into two because ethanol was the carrier solvent for the E2. cryovials for E2 and T analyses) and immediately frozen on liquid nitrogen for later determination of plasma E2 and T concentrations.

277 Parrott et al.

Results

Fish were successfully exposed to model STP effl uent for 21 days. Water temperatures remained relatively constant throughout the exposures (10.3 ± 1.3°C, mean ± standard deviation, n = 130). Stable pH (7.1 ± 0.28, n = 130) and adequate dissolved oxygen (10.3 ± 1.0 mg/L, n = 130) were maintained in all test solutions. Ammonia was signifi cantly higher in the 10% primary model STP effl uent (p < 0.001, ammonia concentration = 1.26 ± 0.19 mg/L, n = 14), compared with all other treatments (that had ammonia concentrations ranging from 0.59 to 0.73 mg/L). During exposures conducted in years 1, 2, and 3 of the study (1997, 1998, and 1999) fi sh survived exposure to 10% primary model-scale sewage effl uent Fig. 1. Mean plasma T concentrations (pg/mL, ± standard (model STP effl uent) and 100% secondary model STP error) of immature rainbow trout exposed to control wa- effl uent, and to high (100 μg/L) and low (25 μg/L) E2. ter (con), 10% primary treated (prim), or 100% secondary There were no treatment-related mortalities over the 21- treated (sec) model-scale STP effl uent for 21 days during day exposures, although some fi sh died when they jumped years 1, 2, and 3 of the study. Asterisks show signifi cant dif- out of exposure pails, and all fi sh in 10% primary model ferences in T concentration compared with control fi sh, with STP effl uent in year 2 died from human curiosity, when p value indicated. night-time janitorial staff removed the air-stone to view fi sh. Exposure of rainbow trout to primary and secondary- There was no dramatic hepatic EROD induction in treated model STP effl uent for 21 days increased plasma any of the treatments (Table 1). Mean ERODs ranged T concentrations (Fig. 1). T was signifi cantly higher in from 1.75 to 5.97 pmol/mg of protein/min (Table 1). fi sh exposed to primary-treated and fi nal sewage effl uent There were two treatment related increases in liver- in year 1. Exposure to 10% primary-treated model STP somatic index (LSI). In year 1, trout exposed to 10% effl uent caused a 4-fold increase in plasma T concentration primary or to 100% secondary-treated model STP (from 190 pg of T/mL of plasma in control fi sh, to 803 effl uent had signifi cantly increased LSIs (2.10 and 1.92, pg/mL in primary effl uent exposed fi sh), while exposure respectively, versus a mean LSI of 1.46 for controls; Table to 100% secondary model STP effl uent caused a 5-fold 1). This pattern was not repeated in years 2 and 3. Trout increase in mean plasma T concentration (from 190 pg LSIs throughout the experiments ranged from 1.18 to of T/mL of plasma to 1,020 pg/mL in secondary effl uent 2.10 (Table 1). exposed fi sh) (Table 1). In year 2, similar trends were seen for fi sh exposed to 100% secondary-treated model STP Discussion effl uent: Fish had signifi cantly increased mean plasma T concentrations (196 pg/mL in control fi sh, 346 pg/mL Ammonia in Model STP Efﬂ uent in secondary effl uent exposed fi sh), although increases were not as dramatic, only 1.7-fold above mean plasma One of the factors that can infl uence fi sh survival and T concentrations of control fi sh, compared with 5-fold responses to MWWEs is ammonia. Ammonia was very above controls the previous year (Table 1). There were high in our 100% model-scale primary-treated effl uent, no data for fi sh exposed to primary-treated model STP about 13 mg/L compared with 0.700 ± 0.28 (n = 23) effl uent in year 2, as all fi sh died due to human error mg/L in the secondary-treated effl uent. Therefore, 10% (see above). In year 3, rainbow trout plasma T was again primary effl uent was the highest concentration of increased with exposure to primary and secondary- primary effl uent that we could expose rainbow trout to. treated model STP effl uent. Increases in mean plasma T The ammonia concentration was 1.26 ± 0.19 (n = 14) concentrations were about 2.5-fold times those of control mg/L in the 10% primary model STP effl uent treatment. fi sh plasma T concentrations (43.6 pg/mL in control fi sh This concentration was about 40% of the 96-h LC50 of versus 109 pg/mL in primary effl uent exposed fi sh and 2.8 mg of NH3 per litre for pH 7.5 water (average LC50 113 pg/mL in secondary effl uent exposed fi sh) (Table 1). for 32 freshwater fi sh species, reviewed in Randall and Exposure to low (25 ng/L) and high (100 ng/L) E2 Tsui 2002). However, two conditions in our experiments caused increases in circulating E2 in blood of trout in would have protected the fi sh somewhat from ammonia years 1 and 2, but not in year 3 (Table 1) when low E2 toxicity: At the lower pH of our exposures (7.1), the + exposure appeared to cause a decrease in circulating E2 ammonia was more in the NH4 form, and less was in plasma. E2 exposures caused no signifi cant changes available in the uncharged NH3 form that crosses in circulating T in trout blood, except for year 3 when membranes. Secondly, the fi sh in our experiments were exposure to high E2 appeared to increase circulating fed every 2 days, and feeding decreases the toxicity of plasma T (Table 1). NH3 to rainbow trout (Wicks and Randall 2002).

278 Model STP Effl uent Affects Sex Steroids in Trout

279 Parrott et al.

Changes in Plasma Steroid Concentrations of Rainbow Trout Exposed to Model STP Efﬂ uent

Exposure of trout to sewage effl uent increased circulating T levels 1.7- to 5-fold over control fi sh. These results are similar to those seen by Tremblay and Van Der Kraak (1999) after exposure of trout to the plant sterol β-sitosterol for 21 days. As well, juvenile rainbow trout exposed to E2, retene, and β-naphthofl avone in our lab have shown increases in plasma T concentrations (data not shown). One study has shown increased T concentrations. Porter and Janz (2003) found MWWE- exposed male longear sunfi sh had increased plasma T compared with fi sh from a reference site. Fig. 2. Mean plasma E2 concentrations (pg/mL, ± standard The increases in T seen with exposure to STP error) of immature rainbow trout exposed to control wa- effl uents were unexpected, as most controlled studies ter (con), 10% primary treated (prim), or 100% secondary of lab fi sh exposed to MWWEs have shown increased treated (sec) model-scale STP effl uent for 21 days during E2 and increased Vtg, and demasculinized fi sh (see years 1, 2, and 3 of the study. Asterisks show signifi cant dif- introduction). It was, however, diffi cult to directly ferences in T concentration compared with control fi sh, with compare steroid responses from fi sh exposed to primary- p value indicated. treated effl uent versus secondary-treated effl uent, as the trout were exposed to 10% primary or 100% secondary fi sh exposures, we could not expose several replicates effl uent (due to differences in toxicity from ammonia, per model STP effl uent at one time. Only one replicate BOD, and TSS in each effl uent). Decreased sex steroid (of 10 fi sh) was tested for each treatment in each year concentrations are usually seen with pulp mill effl uent of the research. A better design would have been three exposures. Studies of fi sh (trout, goldfi sh, mummichog, replicates of 3 or 4 fi sh per treatment per year. However, fathead minnows) exposed to pulp mill effl uents have the logistics of keeping the water temperatures cool for reported decreases in circulating T concentrations the trout made it possible to house only 7 large buckets in (McMaster et al. 1996a, 1996b; Parrott et al. 1999a, the circulated cooling pond baths at one time, and not 21 1999b; Tremblay and Van Der Kraak 1999; Dubé and smaller buckets. Smaller buckets with fewer fi sh would MacLatchy 2000, 2001a, 2001b). have made for better statistical design, but poorer animal E2 was not consistently affected by exposure to husbandry and more stress to the fi sh. In our design, with model STP effl uent (Table 1). In years 1 and 2, there were one large bucket containing 10 fi sh exposed to one model no signifi cant differences in plasma E2 concentrations STP effl uent treatment, the fi sh were pseudoreplicates. among control, primary-treated, and secondary-treated True replication of the exposures occurred over the model STP effl uent exposure groups (Fig. 2). In year 3, three years of the study; fi sh were exposed to primary there was a signifi cant decrease in plasma E2 in trout and secondary-treated model STP effl uent over time exposed to 10% primary-treated model STP effl uent as we returned yearly to assess effects of the model compared with control fi sh plasma E2. Decreased E2 STP effl uents at Sault College. While this was not ideal concentrations are not usually seen with exposure of fi sh statistical design, we feel that this is suffi cient replication to STP effl uents. Rather, T concentrations were decreased to demonstrate the consistent observed trends in steroid (Folmar et al. 1996, 2001), and usually E2 concentrations hormones observed in the trout (Table 1). and Vtg are increased (McArdle et al. 2000; Folmar et al. 2001; Hemming et al. 2001; Todorov et al. 2002; LSI and EROD in Trout Exposed to Model STP Huggett et al. 2003). Exposure to E2 (low, 25 ng/L; high, Efﬂ uent 100 ng/L) increased concentrations of circulating E2 in blood of trout in years 1 and 2 of the experiments. There was an increase in LSI in year 1 fi sh exposed to Because the study was run over three years, there 10% primary and 100% secondary-treated model STP were undoubtedly some differences in the performance of effl uent. Similarly, Weber et al. (2008) saw increased the model-scale STP. The increased potency of the model LSIs of fathead minnows and creek chub (Semotilus STP effl uent during July (year 1) may have been due to atromaculatus) downstream of two Ontario STPs. biological differences in the responses of rainbow trout, There was no signifi cant EROD induction in any of or may have been caused by differences in treatment the MWWE or E2 treatments. There are few reports of and operation of the model-scale STP in the summer MWWEs affecting MFO activities of fi sh. In contrast, compared with years 2 and 3, which were run in the pulp mill effl uents, oil refi nery effl uents, and water from winter months. PAH-contaminated harbours can induce MFO and The study was also statistically pseudoreplicated. increase LSIs in fi sh (Parrott et al. 1999a). Because of the limitations in space and the design of the

280 Model STP Effl uent Affects Sex Steroids in Trout

Conclusions and in vitro biological screening. Environ. Sci. Technol. 32:1549–1558. Exposure of rainbow trout to primary (10%) or Deutsch N. 2000. Evaluation of effl uents from Uppsala secondary (100%) model-scale STP effl uent altered sewage treatment plant with focus on endocrine their sex steroid profi les. Concentrations of T in blood disrupting substances. Study project from the plasma were elevated by exposure to model STP effl uent. Department of Pathology, no. 24, Swedish University Difference in baseline T and E2 concentrations found in of Agricultural Sciences, Uppsala. control fi sh suggest some biological differences in the fi sh Dubé MG, MacLatchy DL. 2000. Endocrine responses used during the three years of the experiments. However, of an estuarine fi sh Fundulus heteroclitus to fi nal despite these differences in control E2 and T, we were effl uent from a bleached kraft pulp mill before and able to detect signifi cant increases in T of fi sh exposed after reverse osmosis treatment of clean condensate. to primary and secondary-treated model STP effl uent Environ. Toxicol. Chem. 19:2788–2796. during the three years of the study. T was signifi cantly Dubé MG, MacLatchy DL. 2001a. Laboratory exposures elevated in blood plasma of fi sh exposed to model-scale of Fundulus heteroclitus to evaluate the signifi cance STP effl uent for 21 days. E2 exposure (positive controls, of reverse osmosis treatment of clean condensates on 25 and 100 ng/L) caused some increased E2 in plasma the endocrine disruption potential of BKPM process of fi sh, indicating exposure and uptake of E2. Few effl uents. Environ. Toxicol. Chem. 20:985–995. changes in plasma T were seen in E2-exposed fi sh. The Dubé MG, MacLatchy DL. 2001b. Identifi cation and exposures provided a quick and simple way to assess the treatment of a waste stream at a bleached-kraft pulp potential steroid disruption caused by primary-treated mill that depresses a sex steroid in the mummichog and secondary-treated model STP effl uent. The rainbow (Fundulus heteroclitus). Environ. Toxicol. Chem. trout steroid bioassay is a useful test for detecting STP- 20:985–995. induced steroid alterations in fi sh. Folmar LC, Denslow ND, Kroll K, Orlando EF, Enblom J, Marcino J, Metcalfe C, Guillette LJ Jr. 2001. Altered Acknowledgments serum sex steroids and vitellogenin induction in walleye (Stizostedion vitreum) collected near a metropolitan sewage treatment plant. Environ. We wish to thank Margo Shaw of the Upper Lakes Contam. Toxicol. 40:392–398. Environmental Research Network, and S. Steller, D. Folmar LC, Denslow ND, Rao V, Chow M, Crain Irwin, S. Livingstone, and other students and staff of DA, Enblom J, Marcino J, Guillette LJ Jr. 1996. Sault College, Sault Ste. Marie, Ontario, for assistance Vitellogenin induction and reduced serum testosterone with the experiments. Funding for the experiment was concentrations in feral male carp (Cyprinus carpio) provided by Environment Canada. captured near a major metropolitan sewage treatment plant. Environ. Health Perspect. 104:1096–1101. Animal Care Hemming JM, Waller WT, Chow MC, Denslow ND, Venables B. 2001. Assessment of the estrogenicity The experiment was conducted under animal care and toxicity of a domestic wastewater effl uent protocols 9703, 9805, 9905, approved by Fisheries and fl owing through a constructed wetland system using Oceans and Environment Canada Joint Animal Care biomarkers in male fathead minnows (Pimephales Committee, Burlington, Ontario. Protocols followed the promelas Rafi nesque 1820). Environ. Toxicol. Chem. guidelines set by the Canadian Council for Animal Care 20:2268–2275. (CCAC 2005). Appropriate Animal Use Data Forms were Hodson PV, Efl er S, Wilson JY, El-Shaarawi A, Maj M, fi led yearly with the Animal Care Committee and with Williams TG. 1996. Measuring the potency of pulp CCAC. mill effl uents for induction of hepatic mixed function oxygenase activity in fi sh. J. Toxicol. Environ. Health References 49:101–128. Huggett DB, Foran CM, Brooks BW, Weston J, Peterson Allen Y, Scott AP, Matthiessen P, Haworth S, Thain JE, B, Marsh KE, La Point TW, Schlenk D. 2003. Feist S. 1999. Survey of estrogenic activity in United Comparison of in vitro and in vivo bioassays for Kingdom estuarine and coastal waters and its effects estrogenicity in effl uent from North American on gonadal development of the fl ounder (Platichthys municipal wastewater facilities. Toxicol. Sci. 72:77– fl esus). Environ. Toxicol. Chem. 18:1791–1800. 83. CCAC (Canadian Council on Animal Care). 2005. Jobling S, Nolan M, Tyler CR, Brighty G, Sumpter JP. Canadian Council on Animal Care Guidelines on: 1998. Widespread sexual disruption in wild fi sh. The care and use of fi sh in research, teaching and Environ. Sci. Technol. 32:2498–2506. testing. ISBN: 0–919087–43–4, p. 94 Larsson DGJ, Adolfsson-Erici M, Parkkonen J, Desbrow C, Routledge EJ, Brighty GC, Sumpter JP, Pettersson M, Berg AH, Olsson P-E, Förlin L. 1999. Waldock M. 1998. Identifi cation of estrogenic Ethinyloestradiol – an undesired fi sh contraceptive? chemicals in STW effl uent. 1. Chemical fractionation Aquat. Toxicol. 45:91–97.

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McArdle M, Elskus A, McElroy A, Larsen B, Benson W, Tilton F, Benson WH, Schlenk D. 2002. Evaluation of Schlenk D. 2000. Estrogenic and CYP1A response of estrogenic activity from a municipal wastewater mummichogs and sunshine bass to sewage effl uent. treatment plant with predominantly domestic input. Mar. Environ. Res. 50:175–179. Aquat. Toxicol. 61:211–224. McMaster ME, Munkittrick KR, Jardine JJ, Robinson Todorov JR, Elskus AA, Schlenk D, Ferguson PL, RD, Van Der Kraak GJ. 1995. Protocol for measuring Brownawell BJ, McElroy AE. 2002. Estrogenic in vitro steroid production by fi sh gonadal tissue. responses of larval sunshine bass (Morone saxatilis x Can. Tech. Report Fish. Aquat. Sci. 1961. Morone chrysops) exposed to New York city sewage McMaster ME, Munkittrick KR, Van Der Kraak GJ. effl uent. Marine Environ. Res. 54:691–695. 1992. Protocol for measuring circulating levels of Tremblay L, Van Der Kraak G. 1999. Comparison gonadal sex steroids in fi sh. Can. Tech. Report Fish. between the effects of the phytosterol β-sitosterol Aquat. Sci. 1836. and pulp and paper effl uents on sexually immature McMaster ME, Munkittrick KR, Van Der Kraak GJ, Flett rainbow trout. Environ. Toxicol. Chem. 18:329– PA, Servos MR. 1996a. Detection of steroid hormone 336. disruptions associated with pulp mill effl uent using Weber LP, Dubé MG, Rickwood CJ, Driedger K, artifi cial exposures of goldfi sh, p. 425–437. In Servos Portt C, Brereton C, Janz DM. 2008. Effects of MR, Munkittrick KR, Carey JH and Van Der Kraak multiple effl uents on resident fi sh from Junction GJ (ed.), Environmental Fate and Effects of Pulp and Creek, Sudbury, Ontario. Ecotox. Environ. Safety Paper Mill Effl uents. St. Lucie Press, Delray Beach, 70(3):433–445. FL, U.S.A. Wicks BJ, Randall DJ. 2002. The effect of feeding and McMaster ME, Van Der Kraak GJ, Munkittrick KR. fasting on ammonia toxicity in juvenile rainbow 1996b. An epidemiological evaluation of the trout, Oncorhynchus mykiss. Aquat. Toxicol. 59:71– biochemical basis for steroid hormonal depressions 82. in fi sh exposed to industrial wastes. J. Great Lakes Res. 22:153–171. Norman A, Örn S, Holbech H, Gessbo Å, Parkkonen J, Received: 1 November 2007; accepted: 11 July 2008. Förlin L, Norrgren L. 2000. Exposure of zebrafi sh (Danio rerio) and rainbow trout (Oncorhynchus mykiss) to 17α–ethinylestradiol and effl uent from a sewage treatment plant, p. 47–62. In Zebrafi sh for Testing Endocrine Disrupting Chemicals, TemaNord 2000:555. Bernan, Lanham, MD, U.S.A. Parrott J, Chong-Kit R, Rokosh D. 1999a. MFO induction in fi sh: A tool to measure environmental exposure, p. 99–122. In Rao S (ed.), Impact Assessment of Hazardous Aquatic Contaminants: Concepts and Approaches. CRC Press Inc., Chelsea, Michigan. Parrott JL, Jardine JJ, Blunt BR, McCarthy LH, McMaster ME, Wood CS, Roberts J, Carey JH. 1999b. Comparing biological responses to mill process changes: A study of steroid concentrations in goldfi sh exposed to effl uent and waste streams from a Canadian bleached sulphite mill. Water Sci. Technol. 40:115–121. Porter CM, Janz DM. 2003. Treated municipal sewage discharge affects multiple levels of biological organization in fi sh. Ecotoxicol. Environ. Saf. 54:199–206. Purdom CE, Hardiman PA, Bye VJ, Eno NC, Tyler CR, Sumpter JP. 1994. Estrogenic effects of effl uents from sewage treatment works. Chem. Ecol. 8:275–285. Randall DJ, Tsui TKN. 2002. Ammonia toxicity in fi sh. Marine Poll. Bull. 45:17–23. Routledge EJ, Sheahan D, Desbrow C, Brighty GC, Waldock M, Sumpter JP. 1998. Identifi cation of estrogenic chemicals in STW effl uent. 2. In vivo responses in trout and roach. Environ. Sci. Technol. 32:1559–1565.

Swimming in Sewage: Indicators of Faecal Waste on Fish in the Saint John Harbour, New Brunswick

Heather A. Loomer,1 Karen A. Kidd,1* Tim Vickers,2 and Alison McAslan1

1Canadian Rivers Institute and Biology Department, University of New Brunswick, Saint John, New Brunswick, Canada E2E 4P1 2Atlantic Coastal Action Program Saint John, Saint John, New Brunswick, Canada E2L 4S4

Despite increased recognition of the risks to the health of humans and the environment, untreated municipal wastewaters are still discharged into waterways worldwide. One of the primary concerns related to its discharge into surface waters is the risk to human health through the transmission of pathogens associated with faecal matter. Saint John, New Brunswick, is one of the few Canadian cities that still releases untreated sewage into its urban waterways and harbour. Water faecal coliform levels, an indicator of faecal waste and associated pathogens, are well above recreational guidelines in some of these areas. Although it is not encouraged by the municipality, recreational fi shing occurs in these areas and this raises concerns regarding the potential for disease transmission during the handling of these fi sh. To investigate the potential for fi sh to be a vehicle of pathogen transmission to humans, the skin of wild fi shes (smelt, Osmerus mordax, and mummichog, Fundulus heteroclitus) and caged mummichog was sampled for faecal coliforms from several sites in Saint John between August and November of 2005. Water faecal coliform levels at sites used for caging studies and wild fi sh collections, and the duration of caging were compared with the number of faecal coliforms on the surface of the fi sh. Skin samples from the two fi sh species collected from the wild indicated elevated levels of fecal coliforms in some locations. Both wild and caged fi sh showed that the amount of faecal coliform on fi sh skin is infl uenced by the water faecal coliform levels.

Key words: sewage, faecal coliforms, fi sh, human health, water quality

Introduction also unknown. If this does occur, the potential exists for pathogen transmission to humans during handling Municipal wastewaters remain a signifi cant threat to of these fi sh, and for the contamination of surfaces and the quality of Canada’s waterways, with untreated or utensils used for their preparation (Fattal et al. 1992; de minimally-treated effl uents being the greatest concern Donno et al. 2002). from both a human and ecosystem health perspective Laboratory studies have shown that the presence of (Chamber et al. 1997; APHA et al. 2005). While there enteric bacteria on fi sh skin is related to the microbial is only a small subset (<5%) of the population whose communities within the water, and that bacterial sewage is released untreated, the presence of combined concentrations in the water and the time of exposure sewer systems and sewage system bypasses during storm are important variables (Fattal et al. 1992; El-Shenawy events can result in the release of additional and likely and El-Samura 1994). Faecal coliforms are a relatively signifi cant volumes of untreated sewage to many of benign group of enteric bacteria which can be used as Canada’s waterways (Chambers et al. 1997). Because indicators of the presence of untreated sewage. The untreated sewage contains numerous human pathogens number of faecal coliforms on fi sh skin was found to including those responsible for hepatitis A, Norwalk fl u, be low (1 colony forming unit [cfu] per 1 cm2) when cholera, and different forms of dysentery, areas receiving exposed to low water concentrations (14 cfu/1 cm3) inputs of untreated sewage are rendered unsafe for (Golas et al. 2002). In another study, a higher density recreational and commercial use (APHA et al. 2005). of faecal coliforms (average 20 cfu/cm2) was observed When untreated sewage is released into waters which on the skin of fi sh exposed to untreated sewage (21,000 support fi sheries and recreational activities, shellfi sh and cfu/100 mL) for 12 hours. However, the bacteria were water quality are monitored for faecal coliforms to assess eliminated from their skin within 48 hours following risks to human health. However, there is less known their placement in coliform-free water (El-Shenawy and about the health risks to humans catching fi sh living in El-Samura 1994). Fattal et al. (1992) found that faecal and moving through these sewage-contaminated areas; it coliform counts on the skin of exposed fi sh reached a is not known whether the skin of fi sh exposed to sewage- maximum of 2.8 cfu/cm2 after 24 hours of exposure derived pathogens can become contaminated and, if so, to 10% wastewater. The corresponding faecal coliform the factors that affect the degree of contamination are concentration at the time of sampling was 290 cfu per 100 mL. Despite these studies, little information is available on the bacterial contamination of wild fi shes * Corresponding author: [email protected] in Canada, or on the factors affecting coliform counts on

283 Loomer et al. fi sh, such as duration and level of exposure, warranting to 3.4(107) in 2005. The Inner Saint John Harbour (two a more detailed investigation on fi sh in waters receiving locations) Indian Town, and Saint’s Rest sites were also inputs of untreated sewage. infl uenced by sewage outfalls, and water faecal coliform At the time of this study, the City of Saint John in New measurements showed moderate levels (230 ± 60, 960 ± Brunswick treated about 60% of its sewage and released 200, and 460 ± 200 cfu per 100 mL). The Tucker Park the remainder untreated into the surrounding waterways and Hazen Creek sites did not receive untreated sewage (City of Saint John 2004). Coliform concentrations up inputs; this is supported by the low water faecal coliform to 3.4x107 cfu per 100 mL have been measured in these levels (35 ± 12 and 34 cfu per 100 mL) measured at these waters (Dupere and Marshall 2005). The main objectives sites. of this study were to determine whether faecal coliforms were present on the surface of wild fi sh caught in waters Sampling Methods receiving untreated sewage, and to investigate the relationships between faecal coliforms on the skin of both Water sampling. As part of a larger monitoring program, wild and caged fi sh and water coliform concentrations. water samples (n = 1 per site per date) were collected for coliform counts, mainly 2 hours before low tide (to reduce Methods dilution by incoming tides). This occurred approximately biweekly at several locations around the Saint John Site Selection Harbour (Table 2) from June to August 2004 and June to October 2005. Samples were collected by moving open, Fish sampling and caging sites were chosen to represent sterile Whirl Pak bags against the current just below the areas of high, moderate, and low sewage inputs to surface of the water. With one exception, water samples estuarine regions within and around the Saint John were collected at a similar location to where the fi sh were Harbour, New Brunswick. These sites were selected caged or collected for the monitoring of wild fi shes. For using both the presence of sewage outfalls in an area, Marsh Creek, routine coliform counts (and water quality water quality parameters (Table 1), and faecal coliform samples) were collected at a freshwater site upstream of counts (Tables 2 and 3) from 2004 and 2005 (see below the location of the caging study and wild fi sh collections. for details on sampling). The Marsh Creek site is at the The water samples were transported to the laboratory mouth of an urban watercourse that receives untreated in a cooler and stored at 4 to 5°C until processed as sewage from eight outfalls before discharging into the described below. Transportation time in the cooler was Saint John Harbour. Sewage inputs were high in this less than 30 minutes, and storage time did not exceed area, and water faecal coliform levels ranged from 1,100 24 hours. Surface water samples (n = 1 per site per date)

284 Indicators of Faecal Waste on Fish

were also collected at the same time for assessment of in Atlantic Canada (Bigelow and Schroeder 2002), were water quality parameters (pH, salinity, total ammonia, caught using short gill net sets (August 23 and 25, 2005) and orthophosphate), and stored on ice until they were at two locations used for recreational fi sheries in the Inner analyzed using standard techniques (APHA et al. 2005). A Saint John Harbour. Fish were sacrifi ced by severing their Corning Model 220 bench type digital pH meter was used spinal cord while they were still trapped in the net, and and calibrated using pH 7.00 and 4.00 buffer solutions. then the fi sh were carefully placed on their side on ice for Total ammonia was measured colourimetrically using transport back to the laboratory. Mummichog (Fundulus the phenate method, and absorbance was determined heteroclitus), a small-bodied fi sh that lives in shallow at 640 nm using a Bausch & Lomb Spectronic 21 estuaries along the eastern coast of North America spectrophotometer. Orthophosphate was quantifi ed using (Bigelow and Schroeder 2002), were collected from the amino acid method. Salinity was determined onsite three locations in the Saint John Harbour area between using a Fisher Scientifi c Accumet portable conductivity September 17 and 30, 2005, using minnow traps and meter with automatic temperature compensation, and beach seines. The fi sh were poured from the traps or converted to equivalent salinity values. net into buckets (previously sterilized with bleach and then rinsed with water from the site) for transport back Wild fi sh. Wild fi sh were caught from four locations to the laboratory (within 20 to 30 minutes of capture). 2 around the greater Saint John area that receive discharges These fi sh were then lethally anaesthetized using CO of untreated sewage (Table 2). Rainbow smelt (Osmerus and sampled within fi ve minutes as described below. All mordax), an anadromous fi sh species commonly fi shed fi sh handling was done in accordance with the Canadian

285 Loomer et al.

Animal Care Committee guidelines and approved by Creek caging site there was an average faecal coliform the Animal Care Committee at the University of New concentration of 8.5(106) ± 3.9(106) cfu per 100 mL (n = Brunswick in Saint John. A subsurface water sample was 15) in 2005 (Table 1). Mummichog (approximately 3- to also collected in a sterile Whirl Pak at each site at the 6-cm in length) were randomly assigned to four cages time of fi sh capture, and faecal coliform concentrations (18 fi sh/cage) that were put into Marsh Creek on August were determined as described below; these samples were 5, 2005, and fi sh were held for up to 8 days. Three fi sh not collected at consistent times of the tidal cycle. per cage were sampled after 12 hours and 1, 2, 4, and 8 From each sacrifi ced fi sh, sterile techniques were days (± 2 hours) as described previously. A subsurface used to remove a 2-cm2 sample of skin tissue and some water sample was taken adjacent to the cage during each underlying muscle with forceps, a scalpel, and a 2-cm2 sampling interval, and these samples were not collected template. Previous studies have shown that the muscle at consistent times within the tidal cycle. tissue of sewage-exposed fi sh contained little or no faecal Results from the fi rst experiment were then used coliforms (Fattal et al. 1992; El- Shafai et al. 2004); for to design a second caging experiment to look at the this reason the presence of muscle tissue in samples was relationship between faecal coliform counts in the water not believed to be a confounding factor in this study. and on the surface of the fi sh. Fish held in the lab were For the smelt, the side that did not rest on the ice was randomly sorted into groups for each cage, and then sampled. To determine the variability of faecal coliform transported as a group into the fi eld in 10-L disposable counts on the surface of the smelt, duplicate or triplicate sterile plastic bags. On day 0, sets of four cages (4 fi sh/ skin samples were taken from six of the fi sh caught in cage) were put in the water at four different locations the Inner Harbour. A fi sh was not used if its side came (Tucker Park, Indian Town, Inner Harbour, and Marsh into contact with any surface during this procedure, or Creek). After four days the cages were removed, and the if the internal viscera were cut during sampling. The fi sh from each cage were put into a new 10-L bag with tissue sample was stored at 4 to 5°C in a sterile 15-mL water from the site for transportation back to the lab. test tube containing 5 mL of sterilized Tryptic Soy Broth Subsurface water samples were also taken on days 0 and (TSB; Difco) prepared according to the manufacturer’s 4 of the experiment from each site. instructions. All of the samples were taken within two hours of the fi sh being caught. Laboratory Methods

Caging experiments. The mummichog used in both Sample preparation. Each skin sample was ground up caging experiments were collected from Hazen Creek in with 5 mL of TSB in a glass homogenizer. The mixture Saint John with a beach seine on July 19 and September was poured into a sterilized test tube (autoclaved for 19, 2005. The fi sh were held for a minimum of two 15 minutes at 240°C) and the homogenizer was rinsed weeks at a consistent temperature (10 to 15°C) and twice with 2.5 mL of magnesium phosphate buffer salinity (15 to 20 ppt), and were fed with 3.0 GR High solution (autoclaved for 15 minutes at 240°C) for a Pro fi sh feed (Corey feed Mills) or Cichlid fl aked food total sample volume of 10 mL. For the wild fi sh and (Nutrafi n) once daily during the warmer months (July to mummichog from the fi rst caging experiment, three 0.05 September), or every second day in the cooler month of to 5 mL subsamples, representing a 1:10 dilution series, October. The fi sh were subsampled (n = 5) before each were put into sterilized graduated cylinders and topped experiment to ensure that they were not contaminated up to 15 mL with buffer. For samples from the second with faecal coliforms before they were put in the cages. caging experiment, three subsamples of 0.07 to 2.5 mL, A second subsample of fi sh (n = 3) was transported into representing a 1:6 dilution series, were taken from the 5 the fi eld and returned to the laboratory to ensure that mL homogenate, put in a sterilized test tube, and diluted contamination did not occur during transport. No faecal to a fi nal volume of approximately 10 mL with buffer. All coliforms were detected in either of these controls for samples were stored at 4 to 5°C for up to 12 hours until both caging studies. they were fi ltered. The cages were made of 4-L plastic cylinders with A sterile standard vacuum fi ltration apparatus was large mesh-covered openings at either end to allow for used to fi lter each homogenate onto a presterilized 47- water fl ow, and were placed at sites subject to tidal mm Millipore fi lter membrane (Fisher Scientifi c) wetted infl uences. At most sites the cages were suspended 0.5 with 20 mL of the magnesium phosphate buffer. The to 1 m below the water surface during low tide, and 3 sample and triplicate rinses of the sample container were m below the water surface during high tide. Cages were fi ltered through the membrane. This same procedure was anchored and arranged on a pulley-type system with used to fi lter the water samples collected from the fi eld ropes to increase ease during sampling and removal of sites, except that the volume of the subsamples varied the cages. from 0.05 to 100 mL based on previous data from those The fi rst caging experiment was conducted at a site sites (Dupere and Marshall 2005). Filter membranes with high faecal coliform counts to determine whether were then placed onto a specialized medium for faecal fi sh skin becomes contaminated with these bacteria, and coliform growth, mFC agar (Difco), that had been poured the importance of exposure time. Upstream of the Marsh into 50 x 11 mm plates approved for water testing

286 Indicators of Faecal Waste on Fish

(Fisher Scientifi c). The agar plate was stored inverted on the bench top until all samples were fi ltered (<1 hour). All plates were then incubated at 44.5 ± 0.5°C for 24 ± 2 hrs to allow the faecal coliform colonies to grow. After the incubation, colonies were counted on the plates using standard guidelines for water sample testing and enumeration (APHA et al. 2005). Blanks using buffer and the presterilized TSB were run every three samples during method development, and then at the beginning and half-way through each subsequent sampling session to ensure that effective sterilization and no cross contamination occurred. If faecal coliform colonies were to develop on the blanks, then all of the samples included in that batch would have been considered to be contaminated. However, this did not occur during either method development or sample analysis. Colonies were randomly selected from eighteen different plates and standard techniques were used to confi rm that they were faecal coliforms, including gram stains, growth on differential media (McKonycase and brilliant green bile broth, Becto-Dickinson), and the presence of oxidase some sites (Tables 2 and 3). At all sites but Marsh Creek, using oxidase test slides (Becto-Dickinson). coliform counts for water samples collected 2 hours before low tide were similar to those collected during fi sh Statistics sampling. For the samples collected at the wild fi sh and caging sites near the mouth of Marsh Creek in August Descriptive statistics were done for each experiment. and September 2005, coliform counts ranged from 1,100

Data from the fi rst caging experiment were log10-plus-1 to >60,000 cfu per 100 mL, and were lower than those transformed and analyzed across sampling times using from the upstream monitoring site (Tables 2 to 4). Water analysis of variance (ANOVA) and Student Newman quality analyses indicated similar pH across sites (except Keuls multiple comparison tests with an alpha value set Hazen Creek), lower salinity at Marsh Creek than at 0.05. The data were log transformed to satisfy the other sites (although these samples were also collected ANOVA’s homogeneity of variance assumption, and 1 upstream of the estuarine caging site), and higher total was added due to the presence of 0 faecal coliforms in ammonia and orthophosphate at sites with higher water some samples. In the wild mummichog sampling and coliform counts (Table 1). the second caging experiment, the amount of faecal coliforms on the fi sh at each site did not follow a normal Wild Fish distribution and the variance was not equal between locations. For this reason, a nonparametric test was used Of the 27 smelt captured in the Inner Saint John Harbour, to evaluate differences between locations. The Kruskal 48% had detectable faecal coliforms on their skin. The Wallis test was used and post hoc comparisons were mean count for these fi sh was 2.5 ± 1.0 cfu/cm2 (Table done using the Mann–Whitney U test. The alpha was 2). Triplicate samples taken from 6 smelt showed that set at 0.05 and type 1 errors were controlled for with bacteria were not uniformly distributed over the surface the Holms Step-down procedure. All statistics were done of the fi sh. The counts on individual fi sh ranged from 0 to using SPSS for Windows version 13.0. Unless otherwise 25 cfu/cm2 and, on 4 of the fi sh, 0 cfu/cm2 was observed indicated, all data are presented as means ± standard in at least 1 of the replicates. error. No statistics were done for the smelt samples. Higher faecal coliform concentrations in the water For reporting purposes, multiple samples from the same were accompanied by higher faecal coliform counts on individual were averaged, and this mean was then used the skin of wild mummichog (Table 2). Coliform counts to generate the overall mean for smelt at this site. on the fi sh from the Marsh Creek site (27 ± 9.4 cfu/ cm2) were signifi cantly higher (p < 0.001 and < 0.001, Results respectively) when compared with the results for fi sh from Hazen Creek (0 cfu/cm2) and Saint’s Rest (0.28 ± Water Samples 0.28 cfu/cm2); counts for fi sh from the latter two sites were not signifi cantly different (p = 0.314; Table 2). Mean faecal coliforms in the water samples ranged from 17 to 4.4(106) cfu per 100 mL, and were highest in Marsh Caging Exposures Creek and lowest in Hazen Creek (Table 2). Across years and within sites, mean coliform counts were similar, but Over the 8 days of the fi rst caging experiment at Marsh there was a considerable range of values within years at Creek, mean faecal coliform counts on the mummichog

287 Loomer et al. ranged from 7.1 ± 3.0 cfu/cm2 at 12 hours to a maximum mummichog is more sedentary and found in and around of 1,200 ± 440 cfu/cm2 on day 4 (Table 4). Water faecal saltwater marshes in brackish waters (Bigelow and coliform concentrations were lowest on day 8 at 20,000, Schroeder 2002). For monitoring programs it is likely and highest on days 1 through 4 at >60,000 cfu/100 mL. more appropriate to use fi sh, such as mummichog, with Faecal coliform counts on the mummichog increased a limited home range in order to reduce uncertainties signifi cantly at each time interval from 12 hours (7.1 related to timeframe of sewage exposure. ± 3.0 cfu/cm2) to 4 days (1,200 ± 440 cfu/cm2) (p < In our caging study, we demonstrated that 0.0001). On day 8, mean faecal coliforms decreased to contamination of fi sh skin with coliforms occurs in as little 14 ± 3.0 cfu/cm2 and were not signifi cantly different from as 12 hours, with maximum contamination observed after what were observed after 12 hours (7.1 ± 3.0 cfu/cm2) 4 days. This contrasts with a previous laboratory study, and on day 1 (21 ± 3.7 cfu/cm2) (p = 0.084 and 0.244, which showed that the maximal concentrations on fi sh respectively; Table 4). skin were reached after a shorter time period (12 hours) The second caging experiment conducted over 4 at lower faecal coliform concentrations, 21,000 cfu/cm2 days at 4 different locations also found higher faecal (El-Shenawy and El-Samura 1994). However, our results coliform counts on fi sh at sites with higher coliform are potentially confounded by the variation in water concentrations in 2005 (Table 3). Mean faecal coliforms faecal coliform counts over this period likely caused by were signifi cantly higher (p = 0.05) on mummichog caged the interactions between the tides and sewage discharges; at Marsh Creek (120 ± 34 cfu/cm2) when compared with although routine water monitoring at these sites was fi sh from the Indian Town site (1.1 ± 0.40 cfu/cm2), and done 2 hours before low tide (Table 2), water samples fi sh from these two sites had coliform levels that were for the caging study and wild fi sh collections were taken signifi cantly higher than those observed on the fi sh caged opportunistically during fi sh sampling. We found that at the Tucker Park (0 cfu/cm2) and Inner Harbour (0 cfu/ faecal coliform counts on mummichog skin were similar cm2) sites (Table 3). The fi sh in three of the four cages put at times when the water concentrations were different. out at the Marsh Creek site appeared discoloured and ill, For example, the faecal coliform levels on the fi sh were and their faecal coliform counts ranged from 48 to 294 not statistically different between 12 hours and 8 days; cfu per 100 mL. In contrast, fi sh in the remaining cage however, the faecal coliform concentrations in water at appeared healthy and their faecal coliform counts ranged these times were >60,000 cfu per 100 mL and 20,000 cfu from 6 to 21 cfu per 100 mL. All other fi sh sampled per 100mL, respectively. These observations could be due in this study did not have any obvious lesions or other to rapid fl uctuations in water faecal coliform levels due external signs of degraded health. to dilution from incoming tides, or to the exceedance of a threshold for sanitary contamination of fi sh skin. Discussion In both the wild and caged fi sh analyzed in this study, the coliform counts were higher on fi sh from sites Although there is considerable focus on the presence of with higher water coliform concentrations, and this faecal coliforms in Canadian waters due to the potential trend was observed for two species of fi sh. In general, co-occurrence of disease-causing pathogens (Chambers these results indicate that fi sh exposed to untreated et al. 1997; APHA et al. 2005), little is known about the sewage in a fi eld setting have sewage derived bacteria on contamination of fi sh with these viruses and bacteria, or their skin, and the level of exposure was indicated by the potential risk to fi shermen when handling fi sh from the quantity of these microorganisms. The presence of surface waters receiving inputs of untreated sewage. This faecal coliforms is not unique to one fi sh species or to study found faecal coliforms on the skin of two estuarine one type of aquatic environment, and has been found on species of fi sh (smelt and mummichog) caught from smelt and mummichog in this study and on tilapia species different locations in the Saint John Harbour area in New exposed to untreated sewage in the laboratory (Fattal et Brunswick. For both of these species, there was a trend al. 1992; El- Shenawy and El- Samura 1994; El-Shafai et towards higher counts of coliforms on individuals caught al. 2004). in locations with higher water coliform concentrations. In addition to coliform abundance in the water, it is This study also found that caged fi sh were contaminated possible that the presence of faecal coliform on fi sh skin with coliform in as little as 12 hours after exposure, and may also be affected by water quality (e.g., pH, salinity, that the abundance of coliform on the skin of these fi sh dissolved O2), species-specifi c differences such as the was related to the level of coliform contamination in quantity of mucus or type of scales, and health of the water collected from the caging sites. fi sh. Water quality measures indicated similar pH (except One of the challenges in studying the relationship Hazen Creek) across sites, but some variability in salinity between exposure to untreated sewage and the microbial (Table 1). However, a systematic examination of the quality of fi sh skin in a fi eld setting comes from the infl uence of water quality on the microbial communities movement of fi sh in and out of these habitats and uncertain associated with the fi shes was not possible given the exposure times. Smelt at this site move throughout different timing of water quality and fi sh sampling; the whole inner and outer harbour area and between nonetheless, it would be challenging in environments like freshwater rivers and saltwater habitats. In contrast, the this one where there are signifi cant tidal infl uences. We

288 Indicators of Faecal Waste on Fish did notice higher coliform counts on fi sh that were visibly organisms on fi sh skin. A risk to human health through unhealthy when compared with healthy individuals from the transmission of diseases during handling of these fi sh the same site, but, because this was only seen at the site may exist and, based on the results of this study, this risk with the highest water concentrations of coliforms, it appears to be greatest in waters receiving the highest was not possible to make any general inferences about inputs of untreated municipal wastewaters as evidenced the relationship between fi sh health and the presence of by water coliform counts. Further study of this area coliform on their skin. would be useful to assess how the microbial communities The use of the cages could have affected the presence associated with fi sh refl ect the microbial communities of faecal coliforms on the mummichog; however, this within the environment and the implications of this for was diffi cult to assess due to the large variation in water human health. faecal coliform levels at the only site where mummichog were both collected and caged, Marsh Creek. When Acknowledgments the results from the second caging experiment and the wild mummichog collected from Marsh Creek were Support for this project was provided by the EJLB compared, the healthier caged mummichog had faecal Foundation, Atlantic Coastal Action Program Saint coliform counts within the same order of magnitude as John, the NSERC Discovery and Canada Research Chair the wild fi sh, indicating that the caging was not likely a programmes, and the University of New Brunswick Saint confounding factor. John Research Fund. We found differences in coliform counts on replicate skin samples from the same smelt. These samples varied References from 0 to 25 cfu/cm2 and indicated that the bacteria were not uniformly distributed over the surface of this species. APHA, AWWA, WEF. 2005. Standard methods for the The 2-cm2 sample represents approximately 2 to 5% examination of water and wastewater. 21st Edition. of the total surface area of a smelt and a much smaller Published jointly by the American Public Health Association, American Water Works Association, and proportion of the entire surface when compared with Water Environment Federation. New York, U.S.A. the same sample from the smaller-bodied mummichog. Bigelow HB, Schroeder WC. 2002. Smelts. Unfortunately, the samples were not differentiated by Order Osmeriformes and Killifi shes. Order location on the fi sh and further analysis of which area on Cyprinodontiformes. In Collette BB and Klein- the fi sh had the greatest occurrence of faecal coliforms MacPhee G (ed.), Fishes of the golf of Maine. 3rd was not possible. Edition. Smithsonian Institution, Washington D.C., Faecal coliforms as a group are relatively benign, U.S.A. although individual members of this group may be Chambers PA, Allard M, Walker SL, Marsalek J, pathogenic (e.g., Escherichia coli 0157). Therefore, Lawrence J, Servos M, Busnarda J, Munger KS, assessing the risk to human health associated with the Adare K, Jefferson C, Kent RA, Wong MP. 1997. presence of faecal coliforms on fi sh is diffi cult. When The Impacts of Municipal Wastewater Effl uents on the sanitary quality of water is tested, a specifi c density Canadian Waters: A Review. Water Qual. Res. J. of faecal coliforms is used to indicate the probability Can. 32:659–713. that a site is highly contaminated with faecal waste City of Saint John. 2004. Wastewater – Environmental and associated disease-causing pathogens are present. Stewardship. Water and wastewater Saint John in However, it is not known what density of faecal coliforms the 21st century. Saint John, NB, Canada. must be present on the fi sh to make similar judgements. de Donno A, Montagna MT, de Rinaldis A, Zonno V, When faecal coliforms were found in high densities on the Gabutti G. 2002. Microbiological parameters in surface of the fi sh, as was found on fi sh from the Marsh brackish water-pond used for extensive and semi- Creek site, it is probable that disease-causing pathogens intensive fi sh-culture: Acquatina. Water Air Soil were also present. However, when faecal coliforms were Pollut. 13:205–214. found in lower densities on fi sh skin, as on smelt caught Dupere C, Marshall S. 2005. Water Quality Monitoring from Long Wharf or on mummichog caged at Indian Program 2005. Atlantic Coastal Action Program Town, the probability that disease-causing pathogens Saint John. Saint John, NB, Canada. were also present is unknown. Further investigation into El-Shafai SA, Gijzen HJ, Nasr FA, El-Gohary FA. the relationships between densities of coliforms and the 2004. Microbial quality of tilapia reared in fecal- presence of enteric pathogens on fi sh skin is required to contaminated ponds. Environ. Res. 95:231–238. address this uncertainty. El-Shenawy MA, El-Samura ME. 1994. Accumulation There is currently very little known about the and elimination of pathogenic bacteria in tilapia fi sh. presence of faecal coliforms on fi sh in waterways Bulletin of the National Institute of Oceanography receiving untreated sewage. This study is one of the few and Fisheries (Egypt) 20:59–68. that has demonstrated the contamination of fi sh skin Fattal B, Dotan A, Tchorsh Y. 1992. Rates of experimental with sewage-derived bacteria and showed that the length microbiological contamination of fi sh exposed to and magnitude of exposure affect the presence of these polluted water. Water Res. 26:1621–1627.

289 Loomer et al.

Golas I, Lewandowska D, Zmyslowska I, Teodorowicz M. 2002. Sanitary and bacteriological studies of water and European catfi sh (Silurus glanis L) during wintering. Archives of Polish Fisheries 10:177–186.

Received: 9 January 2008; accepted: 25 November 2008.

Concentrations of Endotoxins in Waters Around the Island of Montreal, and Treatment Options

Ronald Gehr,1* Santiago Parent Uribe,1 Isabel Fatima Da Silva Baptista,1 and Bruce Mazer2

1 Department of Civil Engineering, McGill University, Montreal, Canada H3A 2K6 2 Department of Pediatrics, McGill University, Montreal, Canada H2X 2P2

Endotoxins are a component of most Gram negative bacteria, and some cyanobacteria. They may be toxic to humans when inhaled or injected, but the effects are unclear when they are ingested. In fact, low concentrations may protect children against certain allergies. Data for endotoxins in Quebec waters are unavailable, hence this study mapped levels in the waters around Montreal, using two commercial test methods. The recently developed factor C method had a greater linear range and was more convenient to use than the widely used Limulus amebocyte lysate (LAL) method. Although the methods gave endotoxin values of the same order, a consistent relationship between the two could not be established. Endotoxin concentrations in the untreated waters varied from 32 to 1,188 EU/mL, comparable in the literature from pristine waters to wastewaters. Values were generally lower in the summer. Filtration is known to be partially effective at removing endotoxins, but the effects of disinfection are not well established. Accordingly, chlorination, ozonation, and ultraviolet light were tested for the destruction of endotoxins in water, at doses found during drinking water disinfection. While chlorine and ultraviolet light had minimal effects on endotoxin levels, ozone could achieve up to 60% reductions at Ct values (concentration × contact time) as low as 2.5 mg·min/L.

Key words: endotoxins, chlorine, ozone, UV

Introduction and Background back to 1973, but their results are orders of magnitude higher than those reported by numerous other authors, The Nature of Endotoxins and may be unreliable. Di Luzio and Friedmann do not mention whether their vials and other glassware were Endotoxins are a constituent of the outer layer of the made endotoxin-free, and their Limulus amebocyte lysate cell wall of most Gram-negative bacteria and some (LAL) test was not as commercialized or standardized as cyanobacteria (Sykora et al. 1980; Anderson et al. 2002). it is today. Several other studies (Jorgensen et al. 1976; They are located in the outermost fi lm of the membrane, Sykora et al. 1980; Burger et al. 1989; Rapala et al. forming part of a greater macromolecular complex called 2002; Rapala et al. 2006) have since been performed to the lipopolysaccharide (LPS) (Anderson et al. 2002). assess the concentrations of endotoxins in waters across Three main components can be identifi ed in these LPS the United States, Denmark, Namibia, South Africa, and compounds: a surface O-specifi c polysaccharide chain, Finland. a core oligosaccharide, and an acylated glycolipid, the lipid A, which anchors the LPS molecule to the outer Methods for Measuring Endotoxins membrane (Stewart et al. 2006). In addition, lipid A is the “endotoxic” (innate immune stimulating) component Several methods have been proposed to identify and of the LPS (Stewart et al. 2006). quantify the amounts of endotoxins. Binding et al. Research has focused mainly on the study of (2004) quantifi ed endotoxins from occupational and the interaction of endotoxins with humans cells by environmental samples by gas chromatography-mass inhalation (Enterline et al. 1985; Gereda et al. 2000; spectrometry (GC-MS) determination of 3-hydroxy fatty Braun-Fahrlander et al. 2002; Smit et al. 2005; Anderson acids (3-OH FAs) present in the lipid A region of the LPS et al. 2007), and little effort has been concentrated on the molecule. However, they recognized that the absolute consequences and absorption mechanisms of endotoxins amount of LPS cannot be determined since LPS from through ingestion (i.e., drinking). different bacterial species will have different 3-OH FAs. The analysis and quantifi cation of endotoxin levels Li et al. (2004) used liquid chromatography (HPLC) in water bodies as well as in drinking water systems is to measure endotoxin levels. They quantifi ed the a relatively recent subject of study. The fi rst published derivatized fatty acids obtained from the lipid component measurements, reported by Di Luzio and Friedmann, date of various LPS molecules. HPLC only measures the physical incidence of OH groups in the lipid A fraction of endotoxins as opposed to the specifi c biological activity of endotoxins which can be quantifi ed by other methods * Corresponding author: [email protected] such as the Limulus Amebocyte Lysate Test.

291 Gehr et al.

Rybka and Gamian (2006) used gas–liquid range of 4 to 17 EU per ng, suggested a conversion factor chromatography–mass spectrometry to detect the levels of 1 ng = 10 EU. of endotoxins but, instead of measuring the OH groups, they measured another component of the LPS molecule, Human Responses to Endotoxins the 3-deoxy-D-manno-2-ocyulosonic acid (Kdo), located between lipid A and the core oligosaccharide. Human response to endotoxin is widely reported in the The authors suggested that Kdo may be considered as a literature, for both Gram-negative and cyanobacterial good candidate for chemical detection of LPS in specifi c endotoxins. Symptoms depend on the exposure route, environments such as human body fl uids. The presence concentration, and exposure intervals. Anderson et of different bacteria in drinking water samples, however, al. (2002) identify the intravenous pathway, air and may render the method unsuitable for such matrices. aerosolized water inhalation, and water ingestion Priano and Battaglini (2005) devised an electro- exposure routes. They cite fever, diarrhoea, vomiting, chemical method to detect endotoxins; it was aimed hypotension, shock, intravascular coagulation, and death at carrying out on-line measurements. The principle is as general symptoms, where the latter only occurs at very based on a gold electrode and a recombinant endotoxin high concentrations. neutralizing protein (ENP). However, LPS is not only Standards are highly variable with regards to absorbed by the ENP, but also by other polymeric matrices acceptable endotoxin exposure levels and they change which leads to possible interferences of proteins. even more depending on the exposure route. For instance, The LAL test has been widely used due to the relative in the case of inhalation, Anderson et al. (2002) quote the simplicity of the equipment needed for the procedure. guidelines proposed by the International Committee on Levin and Bang (1968) discovered that the blood of Occupational Health in 1993: concentrations should not 3 the horseshoe crab, the Limulus polyphemus, contains exceed 200 ng/m to avoid organic toxic dust syndrome, 3 a single cell, the amebocyte, which clots in the presence not exceed 100 ng/m to avoid systemic effects, and be 3 of bacterial endotoxin. Young et al. (1972) then utilized lower than 10 ng/m to avoid airway infl ammation. Heederik and Douwes (1997) refer to the Dutch Expert this fi nding to develop a quantitative relationship. In the Committee on Occupational Standards: for personal chromogenic LAL test developed by Lonza laboratories inhalable dust exposure in an eight hour time-weighted (CAMBREX 2006a), the initial part of the LAL endotoxin average, concentrations should be less than 50 EU/ reaction is used to activate an enzyme which later releases m3 or 4.5 ng/m3; however, in another paper, this same p-nitroaniline (pNA) from the colourless substrate Ac- limit is given as high as 30 ng/m3 (Anderson et al. 2002). Ile-Glu-Ala-Arg-pNA, thus producing a yellow colour Furthermore, Heederik and Douwes (1997) reveal that that can be photometrically measured at 405 to 410 nm. no occupational exposure limit has been established The rate of activation is dependent on the endotoxin anywhere, and they broaden the ‘No Effect Level’ interval concentration. within 170 and 9 ng/m3 (1,700 to 90 EU/m3). Based on The recombinant factor C (rFC) method is based in acute respiratory effects, they establish the ‘No Observed the ability of “factor C” to selectively recognize endotoxin Adverse Effect Level’ (NOAEL) to be 9 ng/m3. The United (CAMBREX 2006b). Nakamura et al. (1986) and Iwanaga States and British Pharmacopoeia have established limits (1993) were able to show a detailed description of the of 0.25 EU/mL for water used for injection. As far as we activation of the Limulus clotting system when induced are aware, no guidelines exist for endotoxins in drinking with LPS. It was found that there are three sequential water. activations of intracellular hemolymph zymogens. A Regarding the ingestion of endotoxins, existing purifi ed and cloned species, the recombinant factor C, is information is merely qualitative in part because the activated by endotoxin binding which then reacts with a mechanisms through which endotoxins are assimilated fl uorogenic substrate to produce a fl uorescent signal. This are not yet fully understood. Snella and Rylander (1977) signal is proportional to the endotoxin concentration in a suggest that there may be a natural defence system at given sample. Fluorescence is measured using excitation/ the intestinal epithelium level, in the form of antibodies emission wavelengths of 380/440 nm at the beginning as that neutralise the endotoxins. They add that ingested well as at the end of the incubation period, and corrected endotoxins should not represent any danger to humans with negative controls. Endotoxin concentrations are unless they have a defi ciency in their immunologic then calculated relative to a standard curve. system. Rapala et al. (2002) note that knowledge on the Endotoxins from Gram-negative bacteria have been occurrence and removal of endotoxins in water samples is found to be as much as 10 times more active than those so limited that guidelines cannot yet be set. Other studies from cyanobacteria (Keleti and Sykora 1982; Anderson on endotoxins in drinking water samples (Jorgensen et et al. 2002; Rapala et al. 2002). Hence the quantifi cation al. 1976; Burger et al. 1989; Rapala et al. 2006) limit of endotoxins in terms of mass (such as ng/mL) has been themselves to the analysis of endotoxin concentrations supplanted by endotoxin units (EU)/mL, where the latter and removal effi ciencies, without relating these levels to is a relative measure based on the activity of endotoxins any standard exposure limit. produced by E. coli bacteria. Anderson et al. (2007), Regardless of the lack of guidelines for maximum having surveyed a variety of articles and recording a concentrations of endotoxins in drinking water, Anderson

292 Endotoxins in Montreal Waters and Treatment Options et al. (2007) have shown that endotoxins can be transferred Removal of Endotoxins During Water Treatment to the vapour phase by showers and humidifi ers. They Processes use literature values which indicate airborne endotoxin health effects at 40 to 53 EU/m3, alternatively symptoms In a fi eld study performed by Burger et al. (1989), it was such as breathlessness, headache, etc. when >40 μg are found that full-scale sand fi ltration reduced endotoxin inhaled over 8 hours or less. Based on these data, and levels by 97.2%; ozonation applied following sand various ranges of exposure times by typical healthy fi ltration decreased the concentrations by an additional individuals to the endotoxin-containing vapours, aqueous 29.7%. Rapala et al. (2002) reported overall reductions concentrations of endotoxins of 1,000 EU/mL and 38,000 at different drinking water treatment plants from 59 EU/mL were used in theoretical equations to predict to 97%, depending on the type and combination of airborne concentrations and exposures. For showers, processes used. Rapala et al. (2006) showed high removal only the extremely high endotoxin concentrations in effi ciencies for sand fi ltration (36 to 96%) and ozonation the water could cause temporary discomfort, but for the (33 to 35%), but no reduction for chlorination. Our own humidifi ers, even the lower concentration could have studies in Montreal have also shown removals during the borderline effects for long exposures (8 h or more). fi ltration stages of approximately 50% at steady-state In contrast, as the incidence of allergic diseases conditions (Parent Uribe 2007). increases, the hygiene hypothesis has generated At the laboratory scale, Anderson et al. (2003a) considerable interest in that certain environmental tested the effect of medium-pressure ultraviolet (UV) exposures have the potential of modifying or educating lamps on endotoxin inactivation. Deionized water was immune responses. Proposed in 1989, the hygiene spiked with 300 and 400 EU/mL, and high UV fl uences hypothesis suggested that in children who had low between 100 and 600 mJ/cm2 were applied, although exposure to bacterial infections, their immune systems practical UV doses for drinking water treatment mature in a manner that is skewed toward developing 2 allergic diseases (Strachan 1989); higher exposures would, normally range between 40 and 100 mJ/cm . With these 2 instead, protect against this development. Epidemiologic parameters, a reduction of 0.55 (EU/mL)/(mJ/cm ) was research has provided more supporting evidence for this reached. The authors claimed that, with this response, up concept. The key elements have been shown to be exposure to 55% removal could be reached if the initial endotoxin to endotoxin (Gereda et al. 2000; Liu and Redmon 2001; concentration was between 50 and 200 EU/mL. Braun-Fahrlander et al. 2002). Furthermore, doses which Anderson et al. (2003b) also tested free chlorine, at can stimulate the immune system are often much lower residual concentrations of 2 and 100 mg/L, and retention than doses that induce symptoms of infection. Hoffmann times of 24 and 210 h, yielding Ct values (concentration et al. (2005) showed that a single exposure to organic dust × contact time) as high as 1.014×106 mg·min/L. The of nonnaïve nonexposed volunteers induced long-lasting reduction rate achieved was 1.4 (EU)/mL·h, which is a symptoms of endotoxin tolerance; Reed and Milton very low value, especially considering that common Ct (2001) found that endotoxins can stimulate immunity values are orders of magnitude smaller. to asthma. Wang and McCusker (2006) showed that A summary of different endotoxin concentrations in exposures of babies to low endotoxin concentrations can raw and drinking water reported in the literature is given assist in preventing certain respiratory diseases. Thus, in Table 1. evidence exists for a potentially benefi cial effect of low dose endotoxin exposure in the prevention of allergic diseases and asthma.

293 Gehr et al.

Objectives Chlorine may interfere with the rFC assay protocol. To avoid this situation, chlorine was quenched with Previous studies have measured endotoxin concentrations sodium thiosulfate before testing for endotoxin. in raw and drinking water systems worldwide, but there The rFC method was preferred to the LAL method has been no study specifi cally for the Montreal area. based on the linearity range of the methods; while LAL Although “high” levels of endotoxins in water are not works in the 0.1 to 1.0 EU/mL range, rFC has a linear considered a health threat, the fact that endotoxins can response between 0.01 and 10 EU/mL. In addition, the be transferred to the vapour phase suggests an interest reagents cost per test was lower for rFC. in such concentration values. Furthermore, the hygiene hypothesis indicates that endotoxins at low concentrations Sampling Protocol can have positive health implications, hence “low” as well as “high” levels are of interest. Therefore an important The fi rst set of water samples was taken on 24th September objective of this work was to obtain background and 22nd October 2005 from 13 freshwater sites in the data on endotoxin levels found in several raw water greater Montreal area. Sites were chosen based on water samples around the Island of Montreal. In doing so, quality data provided by the Réseau de suivi du milieu we also attempted to develop a relationship between aquatique (RSMA), an environmental department of two commercially available methods for endotoxin the City that oversees the quality of surface waters in quantifi cation. Removal during water fi ltration has been Montreal and Laval. reasonably well established in the literature at around Eight sites (BLAP-7, BOU-1.0, FSL-400, JAM-0.0, 50%, but there is little information on the effectiveness LSL-4, MI-4, RDP-140, RDP-550) with high annual of subsequent disinfection processes. Both ozone and UV average fecal coliform counts (≥200 FC per 100 mL) and are becoming more common as primary disinfectants, fi ve sites (FSL-300, IBIZ-11.5, LSL-17, MI-13, RDP-20) with low annual average fecal coliform counts (<200 and chlorine is still widely used. Hence this study also FC/100 mL) were selected. The sampling locations are investigated the effectiveness of these processes, at shown in Fig. 1. All endotoxin tests were performed typical doses used for water disinfection, at removing using the LAL method. An additional sample collection endotoxins. was done in late October 2006. Raw waters feeding three Montreal water treatment plants, called herein A, Materials and Methods B, and C, were also sampled between June, 2006 and January, 2007. The rFC method was used for endotoxin Glassware and Pipettes measurements in this case.

Samples were collected in either 250-mL or 1-L brown UV Inactivation glass bottles with pyrogene-free fl at disk septa. All other glassware was made endotoxin-free by washing with Although Anderson et al. (2003a) tested much higher tap water and rinsing twice with distilled water, then fl uences, fl uences of 40 and 100 mJ/cm2 (mW·s/cm2) baking at 250°C for at least 30 minutes. Pipette tips and were selected herein to be closer to those commonly used microplates were bought pyrogen-free; 1- to 200-μL and for drinking water treatment in Europe and established 100- to 1,000-μL universal fi t pipette tips (Corning Inc, in North American guidelines (Hofmann et al. 2004). Corning, N.Y.) covered all of the pipetting needs. UV irradiation was performed with a collimated beam apparatus. A low vapour-pressure mercury UV lamp Endotoxin Determination emitting at 254 nm was mounted over a collimating tube. UV fl uences were determined following the method The fi rst set of endotoxin measurements (2005) was of Bolton and Linden (2003). UV transmittance of each performed with the QCL-100 chromogenic LAL sample was measured with an Ultrospec 3300 Pro UV/ endpoint assay (Lonza Group Ltd, Walkersville, Md.). visible spectrophotometer (Biochrom, Cambridge, U.K.) The kit included endotoxin standards from the E. coli and the incident intensity of the UV lamp was measured O111:B4 strain. with an IL1400A radiometer and SUL240 probe Subsequently, the Pyrogene recombinant Factor C (International Light, Inc, Newburyport, Mass.). (rFC) endotoxin detection system (Lonza Group Ltd, Water samples from the water treatment plant B were Walkersville, MD), became available. The calibration labelled “raw water” and “fi ltered water,” according to is based on E. coli O55:B5. A FL600 microplate reader their source, and additional spiked samples were created (BioTek Instruments Inc, Winooski, Vt.) equipped with by adding a known amount of endotoxin (E8029-1VL emission/excitation fi lters of wavelengths 360:40 and endotoxin standard, Sigma-Aldrich, Oakville, Ont.) to 460:20, respectively, was used to measure fl uorescence, both raw and fi ltered water samples to yield an initial and the instrument sensitivity was set at 85. Fluorescence endotoxin concentration of about 100 EU/mL. This values were acquired using KC4 microplate reader endotoxin is standardized against USP reference standard software (BioTek Instruments Inc, Winooski, Vt.). endotoxin (RSE), and is extracted from E. coli O55:B5.

294 Endotoxins in Montreal Waters and Treatment Options

Fig. 1. Sampling locations in the greater Montreal area.

295 Gehr et al.

The purpose of spiking was to compare the effect of The complete experiment was performed three times inactivation of high initial endotoxin concentrations vs. in September 2006. Endotoxins were measured by the concentrations in natural waters. Sample volumes of 10 rFC method. mL were poured into 30-mm diameter x 21-mm deep Petri dishes (Kimble glass, Inc. Vineland, N.J.) in duplicate Ozone Inactivation and placed one-by-one beneath the collimated beam for the time necessary to reach the required UV fl uence (3 Endotoxin inactivation by ozone was performed in a minutes for 40 mJ/cm2 and 9 minutes for 100 mJ/cm2). batch reactor with fi ltered water samples in November, Samples were continuously stirred and covered with 2006. Samples were not spiked to keep realistic aluminum foil immediately after the exposure time had conditions, considering that ozonation usually occurs elapsed. Aliquots of each sample were transferred into after fi ltration. 2-mL microcentrifuge tubes, then covered and frozen for Deionized water was saturated with ozone as shown later endotoxin determination. Negative controls were in Fig. 2. Extra dry 99.6% oxygen (MEGS Speciality performed by stirring a sample for the same time but Gases, Inc., Montreal, Que.) was fed into an Ozo 2 VTT without exposing it to UV light. All samples and controls ozone generator (Ozomax, Ltd., Granby, Que.) with a were assayed by the rFC method. nominal production rate of 10 g/hr. Ozone was pumped into a gas washing bottle through a glass-fritted diffuser Free Chlorine Inactivation and transferred into the deionized water until the solution was saturated with a steady-state ozone concentration Tests for endotoxin inactivation by free chlorine were of 3.5 mg/L. The ozone production rate of the ozone performed with fi ltered water samples from the same generator was determined by iodometric titration, as water treatment plant as above. Breakpoint chlorination described by Rakness et al. (1996). It yielded 8.7 g/h. tests were run to establish the chlorine doses needed to Two typical ozone residual concentrations of 0.5 achieve the desired free residual chlorine concentrations. and 1.0 mg/L with two typical retention times of 5 and Samples were not spiked with endotoxins in order to keep 20 minutes were chosen to yield Ct values between realistic conditions, considering that most drinking water 2.5 and 20 (mg·min)/L. The ozone-saturated solution treatment facilities apply chlorination after fi ltration. was added in the amounts described in Table 3 to the Thus the objective was to test the effect of chlorine on corresponding volumes of fi ltered water samples into low endotoxin concentrations. 150-mL Erlenmeyer fl asks. These were previously treated Sodium hypochlorite was used and two different free to render them ozone-demand-free by soaking them in chlorine residual levels, 0.8 and 1.6 mg/L, as well as two a saturated ozone solution overnight, and endotoxin- detention times of 20 minutes and 1 hour were selected, free by the usual method. Sample and ozone-saturated once again to maintain realistic conditions. Tests were solution volumes were chosen so that as much as possible performed in 250-mL graduated Erlenmeyer fl asks which of the fl ask was fi lled to avoid a headspace (ozone were made chlorine-demand- and endotoxin-free by prior degasifi cation). Flasks were covered with endotoxin- exposure to 10 mg/L of chlorine and the usual method free aluminum caps, wrapped in aluminum foil to avoid for endotoxins. Flasks were covered and wrapped with photocatalytical reactions, and mixed in a Junior Orbital endotoxin-free aluminum foil to keep them in the dark to Shaker at approximately 150 rpm for either 5 or 20 avoid photocatalytic reactions, as suggested by Anderson minutes. Duplicates for each Ct value as well as initial et al. (2003b). and ozone-free controls were prepared. Duplicate fl asks were used for all controls, chlorine residuals, and retention times (Table 2). Sample volumes of 50 mL were poured into the fl ask and, at time zero, the appropriate amount of sodium hypochlorite was added to all but two which were used to calculate the initial concentration of endotoxin. Flasks were mixed in a Junior Orbital Shaker (LabLine Instruments, Inc., Melrose Park, Ill.) at approximately 150 rpm for 20 minutes for fl asks one to six, and for 1 hour for fl asks seven to twelve. At the end of each detention time, sodium thiosulfate (0.1 N solution, Fisher Scientifi c, Nepean, Ont.) was added at twice the equimolar requirements to four of the six fl asks in order to quench any remaining chlorine residual; the other two fl asks where no sodium hypochlorite was added served as the oxidant-free controls. Flasks were mixed for another minute to guarantee that all chlorine was quenched, and 1.8-mL aliquots were transferred into 2-mL microcentrifuge tubes that were covered and frozen for later endotoxin concentration determination. Fig. 2. Diagram of ozone production apparatus.

296 Endotoxins in Montreal Waters and Treatment Options

At the end of each retention time, a 5-mL sample would be the method of choice, especially since its use aliquot was withdrawn to measure ozone residuals by involves fewer dilution steps, but more work needs to be the indigo colourimetric method, as described in Section done to compare it with other previously used methods, 4500–Ozone from Standard Methods (APHA et al. especially the LAL. 1998). The absorbance of blank and sample solutions was measured at 600 nm (Hewlett Packard 8452 diode Raw Surface Water Samples and Raw Waters array spectrophotometer). Immediately after the test Feeding the Treatment Plants sample was withdrawn, sodium thiosulfate (0.1 N) was added at twice the equimolar requirements to the The measured endotoxin concentrations in Montreal corresponding fl asks and mixed for 1 minute to ensure surface waters are presented in Fig. 4. that ozone was completely quenched. Sample aliquots of Raw water samples at the inlets of three water 1.8 mL were transferred into 2-mL microcentrifuge tubes treatment plants in the Montreal area were collected which were immediately frozen and stored for endotoxin during the months of June, August and October 2006 concentration determination. and January 2007. Endotoxin concentrations for these Endotoxins were quantifi ed by the rFC method. samples are shown in Fig. 5. The measured endotoxin concentrations in Montreal’s Results and Discussion surface waters varied from 32 to 1,188 EU/mL (Fig. 4), with a mean, median, and standard deviation of 245, 142, Comparison of the Two Endotoxin Measurement and 263 EU/ml, respectively. For samples collected on 23rd Methods September 2005, the measured endotoxin concentrations ranged from 32 to 583 EU/mL, with a mean of 177 EU/ Figures 3a to 3d show comparisons for the LAL and rFC ml and a median of 110 EU/mL. Those collected on the data on the same samples, from the plants and seasons, 22nd October 2005 had concentrations varying from 63 and subject to the sample dilutions shown in the fi gure to 1,188 EU/mL, with a mean and median of 313 and captions. Although the two methods give endotoxin 215 EU/mL, respectively. Endotoxin concentrations were concentrations (in EU/mL) in the same general range, it higher in October for eight samples (BLAP-7, FSL-300, is evident from the equations and R2 values for the trend FSL-400, JAM-0.0, LSL-17, MI-4, RDP-20, RDP-550), lines that there is no consistent relationship. The highest in which a 2- to 15-fold increase was detected. There was correlations are found when the same dilutions are used a 50 to 57% reduction in the endotoxin content in the for both methods, which suggests that experimental three samples (IBIZ-11.5, MI-13, RDP-140) collected in errors may be one reason for the weak relationship. In October, and a 10% decrease in the water at BOU-1.0. the three cases with high or reasonably high R2 values, At LSL-4, the endotoxin concentrations deviated by only the LAL values are higher than the rFC by as much as 2 EU/ml between the two months. 4 times. Endotoxin concentrations measured in most It is understandable that the two methods do not give Montreal waters (53 to 246 EU/mL) were comparable to precisely the same values, as they are based on different concentrations detected in waters polluted by agricultural enzymatic responses to the endotoxin, which itself is and urban activities, but a few contained endotoxin comprised of several active components as described activities similar to those found in pristine or wastewaters. above. Therefore it is not recommended to use the two For instance, the water at LSL-17 and RDP-550 on 23rd methods for assessing the same sample sets. Due to its September 2005 had endotoxin concentrations comparable greater linear range and ease of use, the rFC method to those found in pristine waters. Samples collected at

297 Gehr et al.

Fig. 3. Comparison of endotoxin values (EU/mL) measured by the LAL and rFC methods.

IBIZ-11.5 and RDP-140 on 23rd September, at FSL-400, of wastewater that was discharged at Île de Varennes, an JAM-0.0 and MI-4 on 22nd October, and at BOU-1.0 on island less than 1 km east of the site. both sampling dates reached concentrations detected in With the exception of the sample taken at plant B in municipal wastewaters. The distribution of endotoxin June 2006, all other concentrations in raw waters feeding in Montreal’s surface waters was highly irregular on the Montreal plants ranged between 9 and 30 EU/mL both sampling dates; in contrast to the water quality of (Fig. 5). The high concentration of almost 50 EU/mL for most streams and rivers, endotoxin concentrations were this sample may have been due to contamination of the not lower upstream than downstream. As the results sample during transport or sample handling inside the indicated, concentrations did not increase but fl uctuated lab, since the concentrations for the other samples for as the water moved downstream in all cases. In addition, plant B were lower and similar to those from plants A endotoxin concentrations were generally higher in and C. Overall, there is a tendency for concentrations October than in September. The most noticeable increases to decrease during August and to increase during the in concentration were at JAM-0.0 and FSL-400 where a spring and fall (June and October). Contrary to what 14- and 10-fold increase was observed, respectively. The was expected, concentrations during the winter, when increase in endotoxin activity could have been caused by water temperatures are at their lowest levels, were not fall overturn which brings decomposing algae and other the lowest of the year; in fact, concentrations during endotoxin-inducing substances to the water surface. It the summer were the lowest. Thus, low temperatures do also could have been due to the runoff from the 3-day not necessarily mean low endotoxin levels. One possible heavy rainfall prior to the sampling day in October, explanation for this trend is that spring runoff may be which compromised the water quality. Along with the contributing to the increase in endotoxin concentrations. two aforementioned possibilities, the 10-fold increase in Also, due to the fact that endotoxin can be found free or endotoxin concentration in October at FSL-400 might bound to the cell, although bacteria may have already have been caused by the overturn and lateral dispersion died, endotoxins may still be present in water bodies,

298 Endotoxins in Montreal Waters and Treatment Options

Fig. 4. Endotoxin concentrations at the 13 Montreal surface Fig. 6. Endotoxin inactivation by UV in raw water (n = 2). water sites, measured using the LAL method. The bars represent the mean of duplicate measurements with T lines delimiting half the range between the two measurements.

Fig. 5. Endotoxin concentrations in raw waters, measured using the rFC method (n = 2). irrespective of the low temperatures during the winter. Associated with this may be the reduced microbial activity Fig. 7. Endotoxin inactivation by UV in fi ltered water in cold water, as free endotoxin/LPS may be consumed/ (n = 2). The bars represent the mean of duplicate measure- degraded by other bacteria. This idea is supported by ments with T lines delimiting half the range between the two Sykora et al. (1980); in their study they found that high measurements. bacterial counts did not always occur concurrently with high endotoxin levels in water samples. and Rapala et al. (2002, 2006) (see Table 1). While The differences in the raw water endotoxin the fi rst study found values as high as 1,100 EU/mL, concentrations supplying plants B and C for each concentrations in the last two ranged between 4 and season are not signifi cant, which is logical since both 356 EU/mL. One can observe from these studies that plants obtain their water at the same intake from the St. endotoxin levels in raw water samples are highly variable Lawrence River. On the other hand, plant A takes its raw and mainly dependent on the water source; however, water from Lake St. Louis and, except for the sample even though two different sources were included in this taken in June, concentrations in the remaining three study, endotoxin levels did not vary considerably from samples are slightly higher than those for plants B and C. one source or season to the other. In general, both sources provide water with an endotoxin level within the same range. Endotoxin Inactivation by UV Levels found were generally lower or in the lower part of the interval than those described in the literature. The results for endotoxin inactivation by UV light in raw The three studies that give endotoxin concentrations in and fi ltered water samples can be found, respectively, in raw water samples were those of Burger et al. (1989) Fig. 6 and 7.

299 Gehr et al.

In general, experiments carried out in raw water experiments was 7.6. Results for endotoxin inactivation samples (Fig. 6) showed a higher endotoxin inactivation, by free chlorine are presented in Fig. 8, with each point ranging between 8 and 22%, compared with the fi ltered representing the mean of duplicate measurements and T water (Fig. 7). Nevertheless, there is no clear tendency for bars omitted for clarity. endotoxins to decrease as the UV fl uence increases since, From Fig. 8, we see that initial concentrations in in some cases, there appeared to be a higher endotoxin the samples (controls) ranged between 7 and 10 EU/mL; inactivation at the lower UV fl uence of 40 mW·s/cm2 these were the actual endotoxin concentrations found in with a subsequent increase in the concentration when fi ltered water. Endotoxin removals were within a range UV fl uence was increased to 100 mW·s/cm2. This effect of 11 to 25%. However, similar to the UV experiments, is reproduced irrespective of the sample being spiked increases as high as 8% were also measured, which or unspiked, a situation that suggests that endotoxin suggests that chlorine would have little or no effect on inactivation by UV light is not affected by the initial endotoxin levels. endotoxin concentration in the sample. In a study performed by Anderson et al. (2003b) on Results from the fi ltered water samples (shown in endotoxin inactivation, the effect of several oxidants, Fig. 7) were more variable than those from the raw water including free chlorine, was analyzed. For free chlorine, samples, as shown by the larger T lines obtained. The considerably higher residual doses of 2 and 100 mg/L high endotoxin concentration for the unspiked sample at were applied. Also, much longer retention times of 24, 40 mW·s/cm2 in Fig. 7 is likely an error. 120, and 169 h were used, thus giving Ct values from A similar experiment was performed by Anderson 2,880 to 1.014x106 mg·min/L. As mentioned earlier, et al. (2003a), who obtained higher endotoxin removals the smaller concentrations and retention times chosen overall. However, the authors used much higher UV for the current study were meant to mimic realistic fl uences ranging from 100 to 600 mW·s/cm2, as well as conditions at drinking water treatment facilities. With a medium-pressure UV lamp, compared with the low- their parameters, Anderson et al. were able to achieve pressure one used in the current experiments. Medium- endotoxin inactivation rates of 1.3 and 1.4 EU/(mL·h-1) pressure UV lamps emit light over a broad spectrum, when using, respectively, 2 and 100 mg/L of chlorine including UV as well as visible light, whereas low- residuals. The similar inactivation rates for the two doses pressure UV lamps emit essentially monochromatic UV used suggest that the inactivation rate is independent of light at 254 nm. It is possible that light at wavelengths the chlorine dose, and in fact is virtually zero. other than 254 could have played a role in the endotoxin With the results from Fig. 8, it is diffi cult to establish degradation. In the natural environment sunlight may similar inactivation rates to those obtained by Anderson cause endotoxin degradation, but the specifi c effect et al. (2003b), since values decrease and increase without of different wavelengths is not established at present. following a well-established trend. In their study, Anderson et al. (2003a) developed a linear model to Anderson et al. also concluded that endotoxin inactivation predict theoretical endotoxin concentrations after UV rates were relatively small in terms of signifi cance for light exposure for smaller UV fl uences depending on the drinking water treatment. This idea is confi rmed by initial endotoxin concentration of a given sample. With the results from Fig. 8 where it can be seen that, when their model, a fi nal concentration of 178 EU/mL and used at typical drinking water doses, chlorine is, indeed, 145 EU/mL can be achieved when using, respectively, minimally effective for endotoxin inactivation. fl uences of 40 mW·s/cm2 and 100 mW·s/cm2 in a sample with an initial concentration of 200 EU/mL. These values represent 13 and 28% removal effi ciencies, respectively. In the current experiments with spiked raw water samples, an initial concentration of approximately 200 EU/mL was used. When applying a fl uence of 40 mW·s/ cm2, an endotoxin concentration of 182 EU/mL was achieved, representing a 13% removal; this is consistent with the model proposed by Anderson et al. However, the endotoxin concentration reached with the 100 mW·s/cm2 dose was only 174 EU/mL, or 8% removal, compared with the 28% expected by Anderson et al.’s model. Larger discrepancies were found with this model when attempting to use it with the unspiked raw water sample. Hence, although their model behaved linearly for the 100 to 600 mW·s/cm2 interval, it did not do so for lower UV fl uences (40 to 100 mW·s/cm2).

Endotoxin Inactivation by Chlorine Fig. 8. Endotoxin inactivation by free chlorine (n = 2). Each The mean pH of the samples in the chlorination point represents the mean of duplicate measurements.

300 Endotoxins in Montreal Waters and Treatment Options

Endotoxin Inactivation by Ozone Conclusions

For the ozone results shown in Fig. 9, each point This study showed that surface waters in Montreal represents the mean of duplicate measurements, and bars contained low to moderate amounts of endotoxins when delimit the range. The mean pH of the samples was 7.6. compared with other studies. Endotoxin concentrations Results from Fig. 9 show that endotoxin inactivation varied from 32 to 1,188 EU/mL, were generally higher of up to 57% can be achieved at Ct values of 2.5 mg·min/L, in October than in September, and their distribution in and as much as 74% when increasing the Ct value to waterways was highly random. 20 mg·min/L. Endotoxin concentrations decreased to 3 Endotoxin concentrations in raw water samples to 5 EU/mL and to 2 EU/mL, respectively, for the two were lower than those reported in the literature, and Ct doses used. This is equivalent to a 0.6 log endotoxin they ranged from 9 to 30 EU/mL with a tendency to inactivation with the highest Ct dose applied. decrease during the warmest summer months. Highest As far as we are aware, the effect of ozonation on concentrations were found during the months of June endotoxin inactivation at the laboratory scale has not and October. been published. However, Rapala et al. (2006) reported Laboratory-scale experiments showed that UV light large-scale inactivations of between 33 and 35% for and free chlorine had only a small effect on endotoxin ozonation following coagulation and sand fi ltration. This inactivation at doses typically used in full-scale water lower percentage compared with the laboratory-scale treatment facilities. Inactivation as high as 22% was experiments may be due to the fact that endotoxin was reached with UV light but a consistent reduction previously removed in the coagulation-sand fi ltration could not be established. Similarly, free chlorine was phase. Another factor that may explain this difference is able to inactivate endotoxins by as much as 25%, but the Ct value used at the plant studied by Rapala et al., inconsistently. Ozone proved to be an effi cient method for which unfortunately was not given. endotoxin inactivation, reaching maximum inactivation In a previous study by Rapala et al. (2002), only one levels of 60% with Ct values as low as 2.5 mg·min/L. of the nine treatment plants analyzed included ozonation Thus, endotoxin inactivation is feasible with ozone when as one of the treatment processes. The overall endotoxin reduction for that plant was 96%, placing it as one of applying typical doses used in drinking water treatment the most effi cient ones. As in the 2006 study, the extent facilities. of endotoxin inactivation due to the effect of ozonation was also very low, only 8%. Again, no information was Acknowledgments available on the ozone doses and retention times used. Similarly, Burger at al. (1989) reported a reduction of Funding for this research was provided in part by 30% due to ozonation following sand fi ltration, without an NSERC discovery grant, the Allergen NCE and giving the ozone doses used. CONACYT. The authors would like to thank Maurice Tchio and Laurent Laroche from the City of Montreal and Jean-Luc Beslile from the City of Pointe-Claire for the samples and information supplied as well as Katherine Trajan for her help with the sample collection. We also gratefully acknowledge the valuable input from William Anderson in reviewing the paper.

References

Anderson WB, Dixon DG, Mayfi eld CI. 2007. Estimation of endotoxin inhalation from shower and humidifi er exposure reveals potential risk to human health. Journal of Water and Health 5(4):553–572. Anderson WB, Huck PM, Dixon DG, Mayfi eld CI. 2003a. Endotoxin inactivation in water by using medium- pressure UV lamps. Appl. Environ. Microbiol. 69(5):3002–3004. Anderson WB, Mayfi eld CI, Dixon DG, Huck PM. 2003b. Endotoxin inactivation by selected drinking water treatment oxidants. Water Res. 37(19):4553–4560. Anderson WB, Slawson RM, Mayfi eld CI. 2002. A Fig. 9. Endotoxin inactivation by ozone (n = 2). Each point review of drinking-water-associated endotoxin, represents the mean of duplicate measurements and bars de- including potential routes of human exposure. Can. limit the range. J. Microbiol. 48(7):567–587.

301 Gehr et al.

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Rybka J, Gamian A. 2006. Determination of endotoxin by the measurement of the acetylated methyl glycoside derivative of Kdo with gas-liquid chromatography- mass spectrometry. J. Microbiol. Methods 64(2):171– 184. Smit LAM, Spaan S, Heederik D. 2005. Endotoxin exposure and symptoms in wastewater treatment workers. Am. J. Ind. Med. 48(1):30–39. Snella MC, Rylander R. 1977. The endotoxins: A neglected environmental factor. Sozial- und Praventivmedizin 22(3):137–138. Stewart I, Schluter PJ, Shaw GR. 2006. Cyanobacterial lipopolysaccharides and human health - a review. Environ. Health 5:7. Strachan DP. 1989. Hay-fever, hygiene, and household size. British Medical Journal 299(6710):1259–1260. Sykora JL, Keleti G, Roche R, Volk DR, Kay GP, Burgess RA, Shapiro MA, Lippy EC. 1980. Endotoxins, algae and Limulus amebocyte lysate test in drinking-water. Water Res. 14(7):829–839. Wang Y, McCusker C. 2006. Neonatal exposure with LPS and/or allergen prevents experimental allergic airways disease: Development of tolerance using environmental antigens. J. Allergy Clin. Immunol. 118(1):143–151. Young NS, Levin J, Prenderg RA. 1972. Invertebrate coagulation system activated by endotoxin - Evidence for enzymatic mediation. J. Clin. Investig. 51(7):1790–1797.

Received: 11 October 2007; accepted: 21 October 2008.

303

Infl uence of Polymer Selection on Nutrient Phase Separation for Waste Activated Sludge Thickening at Bench Scale

Allan J. MacDonald1* and Onita D. Basu2

110 – 1441 23rd Avenue SW, Calgary Alberta Canada T2T 0T6 2Carleton University Department of Civil and Environmental Engineering, 1125 Colonel By Drive, Ottawa Ontario Canada K1S 5B6

The use of polymers to condition waste activated sludge prior to thickening is a common practice at domestic wastewater treatment plants. In this study, the performance of various commercially available granular polymers was observed. Thickening by gravity belt thickener was simulated at the bench scale, and the thickened sludge and fi ltrate produced were examined. Laboratory analysis was used to determine the differences in nutrient chemical concentration in the thickened solids and fi ltrate produced by different polymers. By examining the content of nutrient chemicals such as total Kjeldahl nitrogen (TKN), ammonia nitrogen, and total phosphorous, this research showed that polymer choice could affect the chemical composition of thickened sludge and fi ltrate with respect to nutrients. Results showed that the total phosphorous and TKN concentrations in the thickened sludge and fi ltrate were affected by polymer selection, which suggested that the chemical loading on the plant due to fi ltrate recycle, and the nutrient content of thickened sludge delivered to the digester are affected by polymer choice. Polymer optimization studies that examine nutrient properties of fi ltrate and thickened sludge beyond the basic total suspended solids analysis may be advantageous to minimize recycling of these compounds into the fi ltrate phase.

Key words: bench-scale, digestion, nutrient, polymer, sludge, thickening

Introduction biochemical oxygen demand (tBOD5) were examined in the separated thickened solids and liquid fi ltrate to Field research of polymer testing often focuses on the characterize how each polymer performs with respect ability of a given polymer to separate solids from the to the phase separation of nutrients. These nutrients liquid phase within the sludge (Severin and Grethlein may be important to either or both the digestion process 1996; Al-Muzaini and Hamoda 1998; Severin et al. and/or land application of biosolids after digestion. The 1999; Al-Mutairi et al. 2004; Olivier et al. 2004). This importance of this process lies in the fact that the digested information, along with the cost per unit of dry solids is biosolids will be most useful and benefi cial if the digested typically used to select a polymer for sludge thickening. sludge is high in nutrients, since the biosolids can be used In this study, the evaluation of polymers is based not only to supplement the agronomic nutrient requirements of on their performance with respect to solids capture, but plants. also on an evaluation of how different polymers direct In this study, a number of commercially supplied nutrients to solid and liquid phases during thickening. polymers were tested at the bench scale to characterize Thickening by gravity belt thickener (GBT) removes the waste activated sludge (WAS) thickening process the unbound water in the sludge, typically increasing using chemical parameters. Sludge samples were the total solids concentration from about 1% to about collected from a small domestic wastewater treatment 5%, reducing the volume of sludge by about 5 times, plant (WWTP) in Crystal Beach, Ontario. The WWTP and concentrating the sludge before it is introduced at Crystal Beach is an extended aeration facility with no to a subsequent digestion process, (Kiely 1997; Dentel primary clarifi cation, and WAS is thickened by a GBT 2001; Olezskiewicz and Mavinic 2001). Before using a before anaerobic digestion. The design capacity of the GBT to thicken sludge, cationic polymers may be used plant is 9.1 ML/day, with a peak capacity of 27.3 ML/ to neutralize the negatively charged sludge particles to day. The average fl ow through the plant during the study overcome electrostatic repulsion and allow aggregation was approximately 5.5 ML/day, which is about 60% of into fl ocs (Dentel 2001; Chen et al. 2002; Metcalf the design capacity. and Eddy 2003). Polymers tend to be popular for use Understanding the differences between various in gravity drainage systems due to their ease of use polymers with respect to how nutrient chemicals are compared with alternatives such as lime and iron salts directed to both solid and liquid phases during sludge (Novak et al. 1999). Ammonia nitrogen, total Kjeldahl thickening will also allow for tighter control of what is nitrogen (TKN), phosphorous, and total fi ve-day recycled through the WWTP. During thickening, the liquid stream (e.g., fi ltrate from a GBT) is incorporated into the infl ow to the plant. This is an important mechanism * Corresponding author: [email protected] because nutrient chemicals may be removed from

305 MacDonald and Basu wastewater using expensive and/or sensitive processes. and a portion is recycled back to the aeration process. For example, phosphorous is often precipitated out of The remainder of the sludge is wasted from the system activated sludge using coagulants such as ferric chloride. manually in batch fashion, at the discretion of operators. Phosphorous in the form of phosphates can cause WAS is stored in air-mixed tanks prior to thickening on imbalances in the natural aquatic ecosystem, and so a GBT. Thickened sludge then undergoes a mesophilic regulations limit the amount of phosphorous that may anaerobic digestion process in a fl oating roof digester that be present in WWTP effl uent. These negative effects is mixed with a recirculation pump. Digested sludge is have been well documented in closed systems, such as hauled away for disposal at sludge lagoons nearby. Data aquaculture facilities, where nutrient chemicals can from 355 daily sludge samples collected and analyzed at quickly cause imbalances. In fact, some fi eld research has Crystal Beach WWTP are summarized in Table 1. been performed in these applications to use coagulation/ fl occulation and fi ltration to remove nutrient chemicals from recycled water (Cripps and Bergheim 2000; Ebeling et al. 2006). To ensure regulatory compliance, coagulants are mixed with activated sludge in the aeration basins. In the case of ferric chloride, the coagulation mechanism involves a reaction between iron salts and phosphates in the activated sludge, which forms insoluble phosphorous precipitates that settle to the bottom of clarifi ers (Droste 1997). Therefore, if the phosphorous concentration of the activated sludge is increased by the recycling of phosphorous within GBT fi ltrate, the demand for coagulant chemicals will increase to maintain compliance. Understanding how polymers affect the nutrient concentration of solid and liquid streams may prevent Phase 1 Polymer Screening using a polymer that offers solids capture advantages while simultaneously increasing chemical loading through Eight polymers were selected for testing. A bench-scale the plant. Furthermore, in many localities, regulations gravity thickening device was used to simulate the require removal of various forms of nitrogen from plant thickening performance of each polymer. The thickening effl uent. apparatus was made from a two-litre screw-top plastic container with an open bottom. Standard GBT material Materials and Methods (50-micron aperture size) was fi tted over the container and held in place by using waterproof silicone. The belt Sludge Thickening material was allowed to dry in place for approximately 24 hours before testing was started. A custom-made Experimental overview. The bench-scale polymer funnel was screwed on top of the belt material to allow testing process was conducted in two phases. In Phase for WAS mixed with polymer to be tested. 1 (screening), a group of eight samples were used for Sludge was poured into the funnel, and thickened by sludge thickening trials. In Phase 2 (dosage testing), three the gravity belt material. The thickened solids collected polymers were tested at different doses to determine on the surface of the gravity belt material while liquid relative performance over an interval of dosage levels. fi ltrate passed through the porous belt material and was collected in a beaker. Each polymer solution was added to a sample of Polymer preparation for bench-scale testing. Each type WAS at a concentration of 36 mL of polymer (stock of cationic polymer was prepared in a stock solution at concentration 0.1%) per litre of sludge. This corresponds 0.1%, using 20°C tap water, and mixed using a jar test to a dosage of 4.5 g of polymer per kg of sludge solids. apparatus (Phipps and Bird PB700 Standard JarTester) at The dosage selection was chosen to represent the 25 rpm for 1 hour. Polymer solutions were prepared each equivalent average full-scale polymer concentration day for testing. used at the wastewater treatment plant where testing was conducted. The polymer and sludge were mixed Sludge characteristics. The raw wastewater is derived using the same method outlined in Severin and Grethlein from domestic sources with no heavy industrial input. (1996), using 500-mL beakers to pour the mixture back Infl uent wastewater is screened through bar screens, and forth 10 times. Samples (350 mL each) were used to and grit is removed using centrifugal grit chambers, ensure that suffi cient surface area was exposed so that before entering plug fl ow aeration basins with fi ne water could not pool on the surface during drainage bubble diffusers. Ferric chloride is added as a coagulant (Severin et al. 1999). Samples were poured onto the as the activated sludge fl ows from the aeration basins gravity thickener and allowed to drain for 10 minutes. to secondary clarifi ers. Sludge settles in the clarifi ers, Testing was performed as per Standard Methods (AWWA

306 Nutrient Phase Separation During Sludge Thickening

1998) on the separated solid and liquid phases. Solid and wastewater treatment plant where testing was conducted liquid phase testing included: total solids, TKN, total and was considered to be representative of the industry phosphorous, ammonia nitrogen, tBOD5, total suspended standard. solids, and turbidity. The chemical characteristics of the polymers that Results were tested in this study are presented in Table 2. The charge density refers to the percentage of the polymer Polymer Screening (Phase 1) chain that exhibits a positive charge. This percentage may be very low (less than 10 percent), low (10 to 30 Eight polymers were used for the fi rst phase of testing. percent), medium (30 to 50 percent), high (50 to 70 The objective of Phase 1 was to screen polymers for percent), or very high (over 70 percent). The molecular testing in subsequent phases. Figure 1 shows that most weight refers to the average mass in grams of one mole polymers provide thickened sludge of higher total solids of polymer molecules. Among the polymers tested in this content than the control sample, which contained no study, the molecular weight ranged from medium to very polymer. The notable exceptions are polymers 7 and high, characterized as: medium (between approximately 8, which did not improve thickening. Polymers 2 and 5 6 6 10 and 10 g/mol), medium-high (approximately 10 g/ 5 produced thickened sludge with high total solids, 6 7 mol), high (between approximately 10 and 10 g/mol), but with a relatively high variability in the total solids 7 or very high (over 10 g/mol). content between samples (a range of approximately half a percentage point for each polymer), which was Phase 2 Polymer Dosage Selection undesirable. Polymers 1, 4, and 6 produced sludge with relatively high solids content, and were less variable in After Phase 1 polymer testing, three polymers were performance than the others. Polymer 1 at the bench selected to predict appropriate polymer dosages for scale is the same as the polymer used at the full-scale subsequent full-scale thickening. The polymers were GBT application. It is interesting to note that the full- selected for Phase 2 testing based on their ability to scale GBT results were slightly better than the bench- concentrate nutrient chemicals in the solid phase and scale results, indicating some discrepancies with the produce fi ltrate low in nutrient chemicals. Polymer 1 bench-scale testing regime. was automatically selected for dosage testing because it was the incumbent polymer in use at the Crystal Beach WWTP. The dosage range used included the following concentrations (expressed in grams of dry polymer per kilogram of sludge solids): 0, 2.75, 3.65, 4.55, 5.45, and 6.35. The polymer was transferred by pipette to the sludge samples and rapidly mixed for 30 seconds at 300 rpm, then slowly mixed at 25 rpm for 3 minutes on the jar test apparatus. After mixing, 350 mL of fl occulated sludge was poured through the gravity fi lter and allowed to drain for 10 minutes. The thickened solids and the collected fi ltrate were stored separately for testing. A sample of fi ltrate was collected after 12 seconds of drainage, and tested for turbidity. The time selected approximated the contact time between the sludge Fig. 1. Phase 1—Average percent total solids in thickened and polymer mixture at full scale on the GBT at the sludge (n = 2) for each polymer. Dosage: 4.5 g polymer/kg sludge solids.

307 MacDonald and Basu

Fig. 2. Phase 1—Average total phosphorus concentration in Fig. 3. Phase 1—Average total Kjeldahl nitrogen (TKN) fi ltrate (n = 2) for each polymer. Dosage: 4.5 g polymer/kg concentration in thickened sludge (n = 2) for each polymer. sludge solids. Dosage: 4.5 g polymer/kg sludge solids.

The concentration of total phosphorous in fi ltrate Summary of bench-scale polymer screening. The data is shown in Fig. 2. This is a particularly important from bench-scale polymer testing are presented in Table parameter because chemical coagulants are used to 3 (thickened sludge data) and Table 4 (fi ltrate data), remove phosphorous from wastewater. Thus, using a along with the rank of each polymer in each test. In the polymer that reduces the amount of total phosphorous case of the thickened sludge, a rank of 1 indicates the in the recycled fi ltrate can decrease chemical demand and polymer that was most effective at producing sludge with save money. The data suggested that, with the exception the highest total solids percentage or highest nutrient of polymer 7, all polymers reduced the amount of total concentration. In the case of the fi ltrate, the opposite is phosphorous in the fi ltrate compared with control true; a rank of 1 indicates the polymer that was most samples. This phase of testing showed a reduction in effective at producing fi ltrate with the lowest solids fi ltrate phosphorus content of up to 85% with polymers percentage and nutrient concentration. 3 and 4. In a similar study involving an aquacultural Polymers 1, 4, and 6 were selected for further wastewater (Ebeling et al. 2006), a maximum reduction in reactive phosphorus of 41% after using polymers analysis in Phase 2 testing. The polymers were selected was achieved (supporting the fi nding that phosphorous due to their respective solids percentage and low may be directed into the solids phase versus the liquid variability in comparison with the other polymers tested. fi ltrate). In addition, due to the opposing results found with the It is important to discuss the control sample, which total phosphorus and TKN, polymer 6 was selected as was untreated sludge processed through the same testing it displayed the highest total phosphorus in the sludge regime. As can be observed in Fig. 1, the control produced for those polymers with positive cake solids percentage a very low thickened solids percentage, which would be results, and polymer 4 was selected for similar reason expected, however the total phosphorous of this sample with respect to the TKN results. was relatively high compared with the treated samples, Polymer 6 seemed to be most consistent of these with the exception of polymer 7. This result would three, exhibiting lower variability, and producing seem to indicate that the process of polymer thickening thickened sludge with relatively high levels of total generally moves phosphorous into the liquid phase. solids, TKN, and total phosphorous. Both polymer 4 Chen et al. (2006) reported a 10% decrease in fi ltrate and 6 were shown to be relatively effective at producing phosphorous concentration with polymer addition to thickened sludge with consistent solids content, and were sludge, and this value increased upwards of 80% when most effective at producing thickened sludge with high aluminum chloride was used in addition to the polymer. nutrient content relative to other polymer types. From a Examination of the TKN data showed the opposite result, practical perspective, the nutrient content of the sludge which may indicate that polymers generally affect these may be considered a secondary objective compared with nutrients in different ways, an observation which has the solids percentage, because the thickened sludge must been established in previous literature (Westerman and Bicudo 2000). These results highlight the need for more have solid content within the design expectations of full- research on the chemical interactions between polymers scale thickening apparatus, usually between 4 to 6 % and nutrient chemicals. total solids. Figure 3 shows the TKN concentration measured in thickened sludge at the bench scale. Some polymers Bench-scale testing—nutrient values adjusted. In this offer an advantage in terms of concentrating TKN within section, the concentration of total phosphorous and the thickened sludge. For example, polymers 4, 5, and TKN in the solid phase was examined in concert with 6 concentrate approximately twice as much TKN in the percent total solids data by using the total solids thickened sludge relative to polymer 1. percentage to normalize the nutrient data. In other words,

308 Nutrient Phase Separation During Sludge Thickening

the nutrient concentrations observed in this section are a thickened sludge sample, expressed in mg/kg of not expressed in milligram per kilogram of thickened thickened sludge; sludge, they are listed in milligram per kilogram of dry Msol is the mass of the thickened sludge solids, solids. The nutrient concentration data were divided by measured after drying for 24 hours at 105ºC, the total solids percentage in this analysis. The formula expressed in kg; for normalizing the nutrient concentration is shown in Msam is the mass of the thickened sludge sample equation 1: before drying, expressed in kg.

This analysis ties the nutrient concentration of the sludge to the total solids content, thereby correcting for (1) differences in the liquid content of the sludge. In Phase 1 testing, the nutrient concentration of thickened sludge was measured by sampling the thickened sludge as it is Where: thickened by the gravity fi lter. This means that the sludge Cn is the normalized concentration of nutrient n in samples may have different percentages of solid material. thickened sludge, expressed in mg/kg of dried solids; For example, sludge thickened using polymer 5 averaged Cts is the concentration of nutrient n measured in over 5 percent total solids, while sludge thickened using

309 MacDonald and Basu polymer 7 averaged under 1 percent total solids. Since lowest levels of total phosphorous and TKN per kilogram the sludge solids are generally a potent source of nutrient of dried solids on average. This suggests that the polymers chemicals, sludge with a higher solids percentage also may act against the concentration of nutrient chemicals has a higher concentration of nutrient chemicals (Cripps within the sludge to varying degrees. Perhaps polymers and Bergheim 2000). By dividing the nutrient chemical cause chemical reactions that alter nutrient chemicals, concentrations by the percent total solids, it is possible making them unavailable for other chemical reactions. to characterize the chemistry of the sludge to prevent Polymers may cause damage to organisms in the sludge, differences in solids content from affecting the results. causing cell lysis and/or breakdown of organics, which In Fig. 4, the total phosphorus concentration of the could result in higher nutrient concentration in the thickened sludge is examined, while in Fig. 5, the TKN liquid phase. In the future, a more in-depth look at the concentration of the sludge is displayed. interactions between polymers and nutrient chemicals in sludge during thickening may shed some light on the effects of polymer conditioning on the availability of nutrients in thickened sludge. A similar analysis was not performed using the fi ltrate data due to a relatively high level of variability in the solids content of fi ltrate samples.

Polymer Dosage Selection (Phase 2)

The purpose of the Phase 2 testing plan was to determine the relative performance of polymers over a range of dosages (See Materials and Methods section). This testing was designed to determine the optimal dose of each polymer, and also to determine how polymers behave at doses higher or lower than the optimum dose. This analysis can also be useful for predicting the Fig. 4. Phase 1—Normalized average total phosphorus relative advantages between polymers with respect to concentration in thickened sludge (n = 2) for each polymer. solids capture and nutrient phase selection in a full-scale Dosage: 4.5 g polymer/kg sludge solids. application (MacDonald and Basu 2007). Figure 6 shows the total phosphorus concentration of thickened sludge. The data indicated that increasing the dosage past a certain point only caused a marginal improvement in lowering sludge total phosphorus, and that increasing further may actually cause thickened sludge with lower total solids. Similar effects were observed in a sludge thickening study focused on nutrient- rich sludge from an aquaculture facility (Ebeling et al. 2006). This is not surprising because as the polymer is added, the zeta potential of the solids will change and effectively decrease towards neutral and then continue to a higher positive value as more polymer is added, increasing repulsive forces and causing thickened sludge with lower total solids, and thus lower concentrations of nutrient chemicals. At lower doses, polymers 4 and 6 produced sludge Fig. 5. Phase 1—Normalized average total Kjeldahl nitro- with a higher concentration of total phosphorus gen (TKN) concentration in thickened sludge (n = 2) for relative to polymer 1. The data showed that to achieve each polymer. Dosage: 4.5 g polymer/kg sludge solids. a thickened sludge, a total phosphorus concentration of approximately 800 mg/L, a dosage of about 4.375 g of dry polymer per kilogram of sludge solids was needed This analysis showed that there may indeed be using polymer 1, while polymers 4 and 6 produced chemical differences in the sludge produced by different sludge with approximately the same total phosphorus polymers. For example, with the exception of polymer 4, concentration using a lower dose of 3.75 g of dry all polymers produce sludge with lower concentrations polymer per kilogram of sludge solids. Thus, the two trial of total phosphorous and TKN per unit weight of dried polymers (i.e., polymers 4 and 6) appeared to be more solids than the control sample, which was not dosed with effective than the incumbent polymer (i.e., polymer 1) at polymer. In fact, polymer 1 produced sludge with the producing phosphorus-rich sludge at low doses.

310 Nutrient Phase Separation During Sludge Thickening

the fi ltrate of approximately 0.05% is possible with the lowest polymer dose, (3.75 g of dry polymer per kg of sludge solids), and higher doses do not appear to result in signifi cantly lower percent total solids. However,

maximum reductions in tBOD5 concentration occur only at a much higher dose, (5 g of dry polymer per kg of sludge solids).

Fig. 6. Phase 2—Average total phosphorus concentration in thickened sludge (n = 2) for different doses of polymers 1, 4, and 6.

Figure 7 shows the TKN concentration of thickened sludge. Polymers 4 and 6 produced sludge with a higher concentration of TKN relative to polymer 1. The data indicated that to achieve a TKN concentration of Fig. 8. Phase 2—Average percent total solids (n = 3; ±1 approximately 2,500 mg/L, a dosage of about 3.75 g of standard deviation) and total fi ltrate BOD5 concentration dry polymer per kilogram of sludge solids was needed (n = 1) in thickened sludge for polymers 1,4, and 6. using polymer 6, while a dosage of about 4.375 g of polymer per kg of sludge was used with polymer 1 to produce sludge with approximately the same TKN Conclusion concentration. Polymer 4 did not produce sludge with TKN concentration above 1,800 mg/L at any dose. The bench-scale testing identifi ed some distinct advantages in using some polymers to thicken sludge solids with respect to both solids capture and nutrient delivery to the solid phase. For example, Phase 1 of testing indicated that polymer 4 produced sludge solids with 40% more total phosphorous and TKN than polymer 2, with similar solids capture and fi ltrate with 40% less ammonia and 50% less total phosphorous. The results also indicated that most polymers used for thickening will promote the movement of total phosphorus into the liquid phase versus untreated samples; therefore, selection would inherently be based on the polymer that delivers the least amount of total phosphorus into the fi ltrate. Phase 2 testing indicated that polymer choice also affects nutrient delivery to thickened sludge differently than it affects solids capture at different doses. Fig. 7. Phase 2—Average total Kjeldahl nitrogen (TKN) The concentration of tBOD5 in thickened sludge was concentration in thickened sludge (n = 2) for different doses shown to drop signifi cantly only after polymer dose was of polymers 1, 4, and 6. increased to approximately 33% above the point when the total solids percentage of liquid fi ltrate levelled off. Figure 8 shows the liquid phase solids percentage, and Phase 2 testing also showed that differences observed tBOD5 concentration, indicating that the dose of polymer between polymers with respect to producing sludge rich may affect the fi ltrate solids concentration differently in total phosphorus and TKN were signifi cant only at than it affects the tBOD5 concentration. Sources of certain doses. Polymers 4 and 6 produced sludge higher tBOD5 and nutrient chemicals are present in various in total phosphorus than polymer 1 at low doses, but chemical forms which may have different responses to the advantage was diminished at higher doses. In general, thickening, and thus the relationship between solids this study showed that bench-scale testing can be used capture and nutrient phase selection is complex (Jardin to optimize polymer dose and selection with respect to and Popel 1996). Achieving a total solids percentage in nutrient delivery during sludge thickening.

311 MacDonald and Basu

Acknowledgments Metcalf L, Eddy H. 2003. Wastewater engineering: Treatment and reuse. Metcalf & Eddy Inc. 4th edition. The authors would like to acknowledge the role that McGraw-Hill Inc., New York. The Region of Niagara Public Works Department, Novak JT, Agerbæk ML, Sørensen BL, Hansen JA. Water and Wastewater Division played in making this 1999. Conditioning, fi ltering, and expressing waste project possible. Funding for this project was supported activated sludge. J. Environ. Eng. 125(9):816–824. by: Carleton University Department of Civil and Olezskiewicz JA, Mavinic DS. 2001. Wastewater Environmental Engineering The Regional Municipality biosolids: An overview of processing, treatment, and of Niagara. management. Canadian Journal of Environmental Engineering 28:102–114. References Olivier J, Vaxelaire J, Ginisty P. 2004. Gravity drainage of activated sludge: From laboratory experiments Al-Mutairi NZ, Hamoda MF, Al-Ghusain I. 2004. to industrial process. J. Chem. Technol. Biotechnol. Coagulant selection and sludge conditioning in 79:461–467. a slaughterhouse wastewater treatment plant. Severin BF, Grethlein HE. 1996. Laboratory Simulation Bioresour. Technol. 95:115–119. of belt press dewatering: application of the Darcy equation to gravity drainage. Water Environ. Res. Al-Muzaini S, Hamoda MF. 1998. Selection of an effective 68(3):359–369. sludge dewatering system for a small wastewater Severin BF, Nye JV, Kim BJ. 1999. Model and analysis treatment plant. Environ. Int. 25(8):983–990. of belt drainage thickening. J. Environ. Eng. APHA, AWWA, WEF. 1998. Standard methods for the 125(9):807–815. examination of water and wastewater. 20th Edition. Westerman PW, Bicudo JR. 2000. Tangential fl ow Published jointly by the American Public Health separation and chemical enhancement to recover Association, American Water Works Association and swine manure solids, nutrients and metals. Bioresour. Water Environment Federation. New York. Technol. 73(1):1–11. Chen G, Yue PL, Mujumder AS. 2002. Sludge dewatering and drying. Drying Technol. 20(4&5):883–916. Chen SH, Liu JC, Cheng GH, Chang WC. 2006. Conditioning and Dewatering of Phosphorous Rich Received: 7 August 2008; accepted: 9 November 2008. Biological Sludge. Drying Technol. 24:1217–1223. Cripps SJ, Bergheim A. 2000. Solids management and removal for intensive land-based aquaculture production systems. Aquac. Eng. 22:33–56. Dentel SK. 2001. Conditioning, p. 278–310. In Spinosa L and Vesilind PA (ed.), Sludge into biosolids. IWA Publishing, London. Droste RL. 1997. Theory and Practice of Water and Wastewater Treatment. John Wiley and Sons, Toronto. Ebeling JM, Welsh CF, Rishel KL. 2006. Performance evaluation of an inclined belt fi lter using coagulation/ fl occulation aids for the removal of suspended solids and phosphorus from microscreen backwash effl uent. Aquac. Eng. 35:61–77. Jardin N, Popel HJ. 1996. Behavior of waste activated sludge from enhanced biological phosphorus removal during sludge treatment. Water Environ. Res. 68(6):975–983. Kiely P. 1997. Environmental engineering. McGraw-Hill Inc., New York. MacDonald AJ, Basu OD. 2007. Optimization of biosolids treatment through analysis of nutrient availability due to polymer conditioning of waste activated sludge, p. 164–178. In Sludge characteristics, rheology, dewatering, pumping conveying and storage. Proceedings of the 2007 International Water Association Specialist Conference. Moncton, New Brunswick.

Filtration du bleu de méthylène, du chrome hexavalent et de l’acide éthylène diamine tétracétique sur une membrane céramique

d’ultrafi ltration à base de ZnAl2O4-TiO2

Filtration of Methylene Blue, Hexavalent Chromium, and Ethylenediaminetetracetic Acid through an Ultra-Filtration Ceramic

ZnAl2O4-TiO2-based Membrane

El Ghaouti Chahid,1 Hayat Loukili,2 Soufi ane Tahiri,1* Saâd Alami Younssi,2 Abdelhak Majouli,2 et Abderrahman Albizane2

1 Faculté des Sciences d’El Jadida, Département de Chimie, BP 20 El Jadida (24000), Maroc. 2 Laboratoire des Matériaux, Catalyse et Environnement, Faculté des Sciences et Techniques de Mohammedia BP 146 Mohammedia (20650), Maroc.

Dans ce travail, nous avons étudié la rétention du bleu de méthylène (BM), du chrome hexavalent Cr(VI) et de l’acide éthylène diamine tétracétique (EDTA) par une membrane ZnAl2O4-TiO2. Les résultats obtenus ont montré que cette membrane présente une charge résiduelle qui dépend fortement du pH. L’effet de la pression et de la concentration sur le taux de rejet des espèces fi ltrées est étudié. La rétention des espèces ioniques est due à un mécanisme fondé sur les interactions électrostatiques entre les charges portées par la membrane et les ions. Le taux de rejet du BM diminue avec l’augmentation du pH, une situation qui provoque une diminution progressive de la charge positive de la membrane. Dans le cas du Cr(VI) et de l’EDTA, un pH alcalin est favorable pour une bonne rétention car la charge membranaire devient moins positive, le chrome est sous forme de 2- CrO4 et l’EDTA apparaît sous ses formes les plus anioniques. L’effet de la pression sur le taux de rejet des solutés a montré que la contribution convective domine aux fortes pressions et conduit à des rétentions maximales du fait que le fl ux convectif du solvant Jv augmente et vient diluer le perméat.

Mots clés : membrane, ultrafi ltration, bleu de méthylène, chrome hexavalent, EDTA

In this work, the retention of methylene blue (MB), hexavalent chromium Cr(VI) and ethylenediaminetetracetic acid (EDTA) by a membrane ZnAl2O4-TiO2 was studied. The results obtained show that this membrane carries a residual charge which depends strongly on the pH. The effect of pressure and concentration on the retention rate of fi ltered species was investigated. The retention of the ionic species is due to a mechanism based on the electrostatic interactions between the membrane charge and the ions. The retention rate of MB decreases with the increase of pH because the resulting alkanisation causes a progressive decrease of the positive charge of the membrane. In the case of Cr(VI) and EDTA, an alkaline pH is favourable 2- for a good retention because the membrane charge becomes less positive, chromium is in the form of CrO4 and the EDTA appears in its most anionic forms. The effect of pressure on the solute retention rates shows that the contribution due to convection predominates at high pressures and favours maximum retention as the convective fl ux of the solvent Jv increases and causes a dilution of the permeate.

Key words: membrane, ultrafi ltration, methylene blue, Cr(VI), EDTA

Introduction les industries pharmaceutiques et agroalimentaires. Ces procédés, souvent qualifi és de techniques propres, sont Les procédés membranaires ont connu un développement la plupart du temps basés sur une séparation physique important dans de nombreux domaines et sont des différents constituants du milieu et permettent de aujourd’hui utilisés à grande échelle dans plusieurs secteurs travailler à température ambiante (Alami Younssi 1995, industriels tels que le traitement des eaux et des effl uents, 1994; Akbari 2002; Hestekin 1998; Saffaj et al. 2003; la purifi cation chimique, le secteur des biotechnologies, Van Gestel 2002). Les avantages des technologies à membranes par rapport à d’autres procédés physico- chimiques tiennent principalement à la facilité de contrôle du système et au respect de l’environnement * auteur-ressource : t_soufi [email protected] qu’elles permettent. De plus, les procédés à membranes

313 Chahid et al. ne nécessitent pas l’ajout de produits chimiques. L’élaboration des membranes d’ultrafi ltration Les membranes minérales présentent certains supportées par des matériaux argileux, leur caractérisation avantages par rapport à leurs homologues organiques, et leur valorisation dans le traitement des solutions comme leur excellente tenue mécanique, thermique contenant des sels métalliques et des colorants organiques et chimique, une facilité d’utilisation et une grande ont fait l’objet de plusieurs travaux de notre équipe durée de vie (Bouzerara et al. 2006; Burggraf and (Loukili et al. 2006; Saffaj et al. 2005; Saffaj 2004, Saffaj Cot 1996; Khemakhem et al. 2004). Les membranes et al. 2004a, b, c; Saffaj et al. 2003). Dans ce travail, nous céramiques sont de plus en plus utilisées, notamment avons étudié la rétention de bleu de méthylène, du chrome pour la purifi cation de l’eau et le traitement des rejets hexavalent et de l’acide éthylène diamine tétracétique par aqueux salins. Le procédé d’ultrafi ltration est situé une membrane céramique d’ultrafi ltration à base d’oxyde entre la nanofi ltration et la microfi ltration, permettant de titane et d’aluminate de zinc ZnAl2O4/TiO2 (50/50) de séparer les constituants d’un milieu selon leur charge déposée sur une couche de microfi ltration à base d’oxyde et leur taille. Une membrane d’ultrafi ltration peut être de zirconium ZrO2. Les deux couches sont déposées sur considérée comme une collection de pores capillaires un support commercial tubulaire. compris entre 1 et 100 nm de rayon. Le débit (D) du solvant à travers la paroi active sera directement Partie expérimentale proportionnel à la pression hydrostatique du fl uide (ΔP), au rayon moyen des pores de la couche (rp), au nombre Matériels et méthodes de pores (n) et inversement proportionnel à l’épaisseur de la membrane (Δx), conformément à la loi de Darcy: Espèces fi ltrées Notre choix s’est porté sur les agents chimiques suivants : bleu de méthylène (BM), 4 chrome hexavalent Cr(VI) et l’acide éthylène diamine D = (K rp n / Δx) ΔP (1) tétracétique (EDTA). Les caractéristiques des espèces Généralement, l’effi cacité de la membrane est caractérisée étudiées sont regroupées dans le tableau 1. par le seuil de coupure (cut-off) qui peut être défi ni comme

étant la masse molaire M (g/mole) qui correspond à une Membrane ZnAl2O4 - TiO2 rétention pratiquement totale (90 % le plus souvent) d’une macromolécule neutre (effet tamis). Dans le cas La membrane utilisée dans cette étude a une structure d’une membrane chargée, les valeurs des taux de rejet multicouche avec une couche fi ltrante en ZnAl2O4/TiO2 mesurées pour des solutés ioniques ne peuvent pas être (50/50) élaborée par la voie sol-gel : déstabilisation de expliquées uniquement à partir de l’effet de taille, elles solution colloïdale. La couche fi ltrante est déposée sur peuvent être aussi dépendantes de la charge en raison des une couche intermédiaire à base d’oxyde de zirconium interactions électrostatiques membranes- solution (Saffaj ZrO2. Le support, sur lequel sont déposées les deux 2004). couches, est un support en alumine α de géométrie Dans le domaine de l’ultrafi ltration, les solutions tubulaire monocanal, de 150 mm de longueur, de 7 mm traitées contiennent généralement des molécules neutres de diamètre intérieur et 10 mm de diamètre extérieur. Le et des ions. Les phénomènes responsables de la sélectivité support présente une structure asymétrique et composite de la membrane sont liés à des interactions de nature obtenue par la superposition de trois couches d’alumine stérique et/ou électrostatique (effet Donnan), ils sont α de taille de grains différente. L’observation par la dépendants du choix du matériau fi ltrant, de sa porosité, microscopie électronique à balayage (MEB) a révélé du type de soluté choisi et des caractéristiques physico- que la couche fi ltrante a une épaisseur de 1,2 μm. Le chimiques de la solution, notamment du pH et de la force diamètre de pore de la membrane, déterminé à partir ionique. Les interactions électrostatiques sont à l’origine des isothermes d’adsorption–désorption d’azote, est de 4 des fortes sélectivités observées en fi ltration lorsque les nm. Les caractéristiques physiques de la membrane sont espèces sont de taille bien inférieure à celle des pores de regroupées dans le tableau 2. la membrane (Loukili 2006; Saffaj 2004).

314 Rétention de BM, Cr(VI) et EDTA par ultrafi ltration

Charge de la couche ﬁ ltrante ZnAl2O4 - TiO2 Mesure de la perméabilité à l’eau

La surface des oxydes minéraux est constituée de La perméabilité d’une membrane à l’eau est un paramètre groupements hydroxyles (-OH) mais l’hydrogène est important pour défi nir les conditions opératoires de plus au moins mobile et la liaison avec le métal est très fi ltration. La détermination de ce paramètre nécessite polarisée. Ceci confère un caractère amphotère aux l’étude de la variation du fl ux de perméation à l’eau en oxydes minéraux. En fonction du pH du milieu, on peut fonction du temps. La perméabilité de la membrane est avoir les réactions suivantes : ensuite obtenue à partir de la pente de la courbe fl ux =

f(ΔP). Le fl ux volumétrique du solvant Jv, le coeffi cient de + + perméabilité L et la pression de travail ΔP sont liés par M-OH (surface) + H (aq) [M-OH2] (surface) p - - la relation suivante : M-OH (surface) + OH (aq) MO (surface) + H2O

J = L . ΔP Ces réactions montrent donc qu’on a un excès de charges v p (2) positives ou négatives, suivant le pH à la surface du solide (Randon 1991). Tests de ﬁ ltration Des travaux antérieurs (Loukili 2006; El Marraki 2001; Saffaj 2004) montrent que la membrane en Une membrane est caractérisée généralement par le taux de rejet R (appelé aussi taux de rétention) de l’espèce ZnAl2O4 - TiO2 (50/50) est chargée positivement pour des pH inférieurs à 9 en présence du sel simple (indifférent) fi ltrée, selon la formule : NaCl. (3) R (%) = (1 – Cp/Ca) x 100 Pilote de ﬁ ltration

où Ca est la concentration initiale de l’espèce dans la

Le montage utilisé est réalisé au laboratoire et il est solution d’alimentation et Cp la concentration de l’espèce fait entièrement en acier inoxydable afi n d’éviter les dans le perméat. problèmes de corrosion. La capacité du réservoir L’effet de la concentration de l’espèce fi ltrée, du pH d’alimentation est de 2 litres et la pression de travail et de la pression sur le taux de rejet est étudié pour les peut être variée jusqu’à 10 bars à l’aide d’une bouteille trois produits. Le pH des solutions fi ltrées est ajusté par d’azote. Le système fonctionne en mode discontinu, la ajout de HCl ou NaOH sans variation du volume initial circulation du liquide est assurée au moyen d’une pompe des solutions. avec une vitesse de 2,6 m/s et le carter permet de recevoir des membranes tubulaires de 15 cm de longueur et 7 Analyse des solutions ﬁ ltrées mm de diamètre intérieur (fi gure 1). La température de fi ltration est maintenue à environ 202 °C à l’aide d’un Le dosage du bleu de méthylène (BM) et du chrome circuit de refroidissement. hexavalent Cr(VI) complexé par la 1,5-diphénylcarbazide La membrane est conditionnée dans l’eau pure (DPC) est réalisé à l’aide d’un spectrophotomètre UV- pendant les 24 heures qui précèdent la fi ltration afi n visible de type (UNICAM UV/VIS) en utilisant des cellules d’obtenir rapidement la stabilité du fl ux de perméat. de quartz. Toutes les mesures sont effectuées à la longueur d’onde correspondant à l’absorbance maximale (λmax) de chaque espèce. La longueur d’onde maximale utilisée pour le dosage du BM est 664 nm. La concentration du chrome hexavalent est déterminée selon la norme internationale (ISO11083 1994) qui repose sur la réaction du chrome hexavalent avec la 1,5-diphénylcarbazide. L’absorbance du complexe « chrome-1,5-diphénylcarbazone » ainsi formé de coloration rouge-violette est mesurée à une longueur d’onde voisine de 540 nm. L’acide éthylène diamine tétracétique (EDTA) est dosé par complexométrie en utilisant une solution de cuivre (10-3M) en présence d’un tampon pH=5 et de l’indicateur coloré PAN (White Figure 1. Schéma du pilote d’ultrafi ltration. 1975).

315 Chahid et al.

Résultats et discussion

Mesure de la perméabilité à l’eau

A partir de la courbe fl ux = f(ΔP) (fi gure 2), nous avons pu déterminer la perméabilité de la membrane comme étant égale à 9,8 L/h·m2·bar.

Figure 3. Effet de la pression sur la rétention du BM (C= 50 mg/L, pH=4,8).

le soluté pour atteindre la membrane de fi ltration. Nous avons étudié également l’effet du pH sur la rétention du BM en utilisant une solution à 50 mg/L. La fi ltration est réalisée pour différents pH : 2, 4, 6, 8 et Figure 2. Variation du fl ux de perméation à l’eau en fonc- 10 (fi gure 6) sous une pression de 8 bars. Les résultats tion de la pression. obtenus montrent que le taux de rejet du BM diminue avec l’augmentation du pH, il est de 86 % à pH 2 et de Filtration du bleu de méthylène 24 % à pH 9. Ceci peut être expliqué par la variation de la charge membranaire avec le pH. Cette charge atteint Les colorants de l’industrie du textile posent de sérieux son maximum à pH acide et diminue progressivement problèmes environnementaux en raison de leur stabilité avec la basifi cation. L’augmentation du pH entraîne une et leur faible biodégradabilité. Le traitement des rejets diminution de la force d’interaction entre la surface de la des industries du textile est donc nécessaire avant leur membrane et le cation du bleu de méthylène. évacuation dans le réseau d’assainissement. La fi ltration des colorants anioniques sur une membrane Cordiérite/ Filtration du chrome hexavalent

ZrO2/ZnAl2O4-TiO2 a été étudiée par Saffaj et al. (2004). Le bleu de méthylène, qui fait partie de la classe des Dans des travaux antérieurs (Saffaj 2004), il a été colorants thiazine, est choisi dans ce travail comme démontré que la couche membranaire ZnAl2O4-TiO2 est substance modèle des colorants cationiques. très effi cace pour la rétention des métaux lourds (chrome Pour étudier l’effet de la pression sur la rétention de ce colorant, nous avons fi ltré des solutions de 50 mg/L à 4, 6, 8 et 10 bars. Les résultats obtenus montrent que l’augmentation de la pression favorise la rétention du soluté (fi gure 3). Par exemple le taux de rejet après 2 heures de fi ltration augmente de 35 à 100 % lorsque la pression passe de 4 à 10 bars. Ceci est dû au phénomène de convection qui est important aux fortes pressions. Sous une pression de 8 bars, nous avons étudié l’effet de la concentration du BM sur son taux de rejet R(%) en utilisant différentes concentrations (25, 50 et 100 mg/L). La fi gure 4 montre une augmentation du taux de rejet avec la concentration. L’augmentation de R(%) est due à la réduction du diamètre des pores de la membrane, favorisant ainsi le rejet du soluté. La diminution du fl ux à fortes concentrations (fi gure 5) peut être expliquée par la formation d’une couche de gel et une accumulation de soluté créant un fl ux de diffusion opposé au fl ux de convection. La couche de gel formée se comporte comme Figure 4. Évolution du taux de rejet en fonction du temps à une nouvelle membrane au travers laquelle devra diffuser différentes concentrations du BM (ΔP= 8 bars, pH=4,8).

316 Rétention de BM, Cr(VI) et EDTA par ultrafi ltration

Cr3+, cadmium Cd2+ et plomb Pb2+), puisque la membrane chargée positivement entraîne une forte exclusion des ions métalliques divalents et trivalents par effet Donnan. Dans ce travail, nous avons étudié la possibilité d’utiliser

la membrane ZnAl2O4-TiO2 pour la rétention du chrome hexavalent Cr(VI). La forme hexavalente du chrome est mise à l’étude, compte tenu de sa grande toxicité. Pour étudier la nature des interactions entre la membrane et les ions, nous avons fi ltré différentes solutions à 25 mg/L sous une pression de 8 bars à différents pH. Comme on peut le constater, le pH a un effet très important sur la rétention du chrome hexavalent (fi gure 7). Le taux de rejet de Cr(VI) est nul à pH=4,93 et maximal (100 %) à pH=10,7. La rétention atteint son maximum à 2- pH très basique car les ions CrO4 sont exclus fortement par la membrane qui devient négativement chargée Figure 5. Évolution du fl ux en fonction du temps à différ- (MO-) dans un milieu alcalin. À pH acide, la membrane entes concentrations du BM (ΔP= 8 bars, pH=4,8). + est positivement chargée (MOH2 ) et les répulsions 2- entre la surface fi ltrante et les anions Cr2O7 sont par conséquent négligeables. L’effet de la pression sur la rétention du chrome hexavalent à pH 9 montre que la valeur maximale de la rétention est atteinte à partir d’une pression de 6 bars. Le taux de rejet ne dépasse pas 50 % à 4 bars; par contre, pour des pressions supérieures à 6 bars, la rétention est supérieure à 90 % (fi gure 8). La contribution convective domine aux fortes pressions et conduit à des rétentions

maximales du fait que le fl ux convectif du solvant Jv augmente et vient diluer le perméat. Pour étudier l’effet de la concentration de Cr(VI) sur le taux de rejet, nous avons fi ltré à pH 9 et sous une pression de 8 bars trois solutions à différentes teneurs en soluté (10, 25 et 50 mg/L). Les résultats obtenus (fi gures 9 et 10) montrent une légère diminution du taux de rejet et du fl ux avec l’augmentation de la concentration du chrome hexavalent. La diminution du taux de rejet lorsque la concentration augmente de 10 à 50 mg/L peut Figure 6. Évolution de la rétention du BM en fonction du être expliquée par l’effet d’écrantage des charges dû à temps à différents pH (ΔP= 8 bars, C=50 mg/L). l’augmentation de la concentration du contre-ion. Les

Figure 7. Évolution de la rétention de Cr(VI) en fonction du Figure 8. Évolution de la rétention de Cr(VI) en fonction du temps à différents pH (ΔP= 8 bars, C=25 mg/L). temps à différentes pressions (C=25 mg/L, pH=9).

317 Chahid et al.

charges de la membrane sont compensées par une partie des contre-ions. L’affaiblissement de la densité de charge surfacique qui en résulte fait chuter la rétention. Dans le cas du chrome trivalent Cr(III) et dans une plage de pH où la membrane est positivement chargée, Saffaj (2004) a montré le même comportement : la

fi ltration de Cr(NO3)3 à pH=3 a révélé une augmentation du taux de rejet de 80 à 96 % lorsque la concentration diminue de 10-2 à 10-3 M.

Filtration de l’acide éthylène diamine tétracétique

EDTA est l’abréviation de « l’acide éthylène diamine tétraacétique ». La formule chimique de cet acide

aminopolycarboxylique est C10H16N2O2. Ce tétra acide est un agent chélatant (ou complexant) capable de former Figure 9. Évolution de la rétention de Cr(VI) en fonction du des complexes métalliques très stables. Il est susceptible temps à différentes concentrations (ΔP= 8 bars, pH=9). d’être utilisé dans de nombreuses applications. En chimie, l’EDTA est utilisé pour traiter des eaux (par exemple dans les lessives), pour éviter les précipitations (entartrage) ou pour doser par complexation les ions métalliques en solution. L’acide éthylène diamine tétracétique présente quatre pKa : 10,3; 6,3; 2,7 et 2 qui correspondent 4- 3- 3- 2- respectivement aux couples : Y /HY , HY /H2Y , 2- - - H2Y /H3Y et H3Y/H4Y. L’EDTA est dangereux pour l’environnement; par exemple, ce produit est capable de libérer des métaux lourds renfermés dans des sédiments. Pour étudier l’effet de la pression sur la rétention de l’EDTA (10-3 M), nous avons fi ltré des solutions à différentes pressions (6, 8 et 10 bars) et à pH normal de la solution préparée (pH=4,8). Comme dans le cas du bleu de méthylène et du chrome hexavalent, nous avons montré une augmentation du taux de rejet de l’EDTA avec la pression (fi gure 11). Nous avons ensuite étudié l’effet du pH sur le taux Figure 10. Évolution du fl ux en fonction du temps à différ- de rejet de l’EDTA. La fi ltration est réalisée à différents entes concentrations de Cr(VI) (ΔP= 8 bars, pH=9). pH (2,5; 4,8; 6,3; 8 et 10,11) sous une pression de 8 bars. Les résultats obtenus (fi gure 12) montrent que le taux de rejet de l’espèce fi ltrée dépend fortement du pH, l’augmentation de ce paramètre permettant une bonne rétention de l’EDTA. Ce comportement peut être expliqué par la diminution de la charge positive de la membrane et par l’apparition des formes plus anioniques de l’EDTA à pH élevé. Après l’étude de l’effet du pH et de la pression, nous avons étudié l’effet de la concentration sur la rétention de l’EDTA et le fl ux en utilisant trois solutions de différentes teneurs en EDTA (5·10-4, 10-3 et 5·10-3M). La fi ltration est réalisée à pH normal de la solution (pH=4,8) et sous une pression de 8 bars. Comme dans le cas du bleu de méthylène, les résultats obtenus ont montré une augmentation du taux de rejet du soluté (fi gure 13) et une diminution du fl ux (fi gure 14) avec l’augmentation de la concentration. Par exemple après 60 minutes de fi ltration, la rétention de l’EDTA passe de 20 % à 44 % et le fl ux diminue de 104,5 à 84,5 L/h·m2 lorsque la concentration Figure 11. Évolution de la rétention d’EDTA en fonction du augmente de 5·10-4 à 5·10-3M. temps à différentes pressions (C=10-3M, pH=4,8).

318 Rétention de BM, Cr(VI) et EDTA par ultrafi ltration

Conclusion

La fi ltration du bleu de méthylène, du chrome hexavalent et de l’acide éthylène diamine tétracétique sur la membrane

d’ultrafi ltration ZnAl2O4-TiO2 (50/50) a montré que le taux de rejet des espèces dépend fortement de la charge de la membrane et du pH. L’augmentation du pH est défavorable pour la rétention du bleu de méthylène et favorable pour la rétention de Cr(VI) et de l’EDTA. Les résultats ont montré également que le fl ux et le taux de rejet dépendent de la concentration de la solution d’alimentation. L’augmentation de la pression du travail favorise la rétention des solutés fi ltrés, ceci étant dû au phénomène de convection qui est importante aux fortes pressions. Figure 12. Évolution de la rétention d’EDTA en fonction du La membrane utilisée a montré des résultats temps à différents pH (ΔP= 8 bars, C=10-3M). intéressants qui permettent d’envisager son utilisation pour le traitement fi nal des effl uents industriels pollués par les colorants et les métaux. Le couplage des procédés physico-chimiques et biologiques de traitement avec la fi ltration membranaire est l’un des objectifs de nos travaux en cours.

Références

Akbari A, Remigy JC, Aptel P. 2002. Treatment of textile dye efl uent using a polyamide based nanofi ltration 5·10-4M 10-3M membrane. Chem. Eng. Process 41:601–609. 5·10-3M Alami Younssi S, Larbot A, Persin M, Sarrazin J, Cot L. 1994. Gamma alumina nanofi ltration membranes. Application to the rejection of metallic cation. J. Membr. Sci. 91:87. Alami Younssi S, Larbot A, Persin M, Sarrazin J, Cot L. 1995. Rejection of mineral salt on a gamma Figure 13. Évolution de la rétention en fonction du temps à alumina nanofi ltration membrane. Application to différentes concentrations d’ EDTA (ΔP= 8 bars, pH=4,8). environmental processes. J. Membr. Sci. 102:123. Bouzerara F, Harabi A, Achan S, Larbot A. 2006. Porous ceramic supports for membranes prepared from kaolin and doloma mixtures. Journal of the European Ceramic Society 26(9):1663–1671. Burggraf AJ, Cot L. 1996. Fundamentals of Inorganic Membranes Science and Technology. Membrane Science and Technology Series 4, Elsevier, Amsterdam, 585. El Marraki Y. 2001. Élaboration et caractérisation de 5·10-4M 10-3M membranes à base d’oxyde de titane et d’aluminate 5·10-3M de zinc pour la nanofi ltration et l’ultrafi ltration, Thèse de Doctorat. Université de Montpellier II, France. Hestekin JA, Bhattacharrya D, Sikdarand SK, Kim BM. 1998. Membranes for treatment of hazardous wastewater. In Meyers RA (ed.), Encyclopedia of Environmental Analysis and Premeditation. John Wiley and Sons Inc. ISO 11083. 1994. Qualité de l’eau – Dosage du chrome Figure 14. Évolution du fl ux en fonction du temps à différ- (VI) – Méthode par spectrométrie d’absorption entes concentrations de EDTA (ΔP= 8 bars, pH=4,8). moléculaire avec la 1,5-diphénylcarbazide.

319 Chahid et al.

Khemakhem S, Ben Amar R, Ben Hassen R, Larbot White WW, Murphy PJ. 1975. Use of Copper-PAN in the A, Medhious M, Ben Salah A, Cot L. 2004. New Selective Titrimetric Assay of EDTA and Its Alkali ceramic membranes for tangential waste – water Salts. Analytical Chemistry 47(12):2054. fi ltration. Desalination 167:19–22. Loukili H. 2006. Filtration des solutions ioniques sur

membranes d’ultrafi ltration en ZnAl2O4/TiO2 Reçu : 19 septembre 2006; accepté : 20 août 2008. déposées sur supports à base d’argiles, Thèse de Doctorat. Faculté des Sciences et Techniques de Mohammedia, Maroc. Loukili H, Alami Younssi S, Ouammou M, Albizane A, Persin M, Larbot A. 2006. Filtration de solutions salines et de colorants sur membranes d’ultrafi ltration

en ZnAl2O4-TiO2 déposées sur support d’argile. Récents progrès en génie des procédés 93 ISBN 2-910239- 67-5. Randon J. 1991. Infl uence de l’interface oxyde-solution sur les performances des membranes d’ultrafi ltration en zircone, Thèse de Doctorat. Université de Montpellier II, France. Saffaj N. 2004. Préparation de membranes d’ultrafi ltration à base d’oxyde de titane et d’aluminate de zinc sur supports d’argiles. Application à la fi ltration de sels métalliques et des colorants. Thèse de Doctorat. Faculté des Sciences et Techniques de Mohammedia, Maroc. Saffaj N, Alami Younssi S, Albizane A, Messaoudi A, Bouhria M, Persin M, Cretin M, Larbot A. 2004a.

Elaboration and properties of TiO2-ZnAl2O4 ultrafi ltration membranes deposited on cordierite support. Sep. Purif. Technol. 36:107–114. Saffaj N, Alami Younssi S, Albizane A, Messaoudi A, Bouhria M, Persin M, Larbot A. 2004b. Preparation and characterisation of ultrafi ltration membranes for toxic removal from wastewater. Desalination 168:259–263. Saffaj N, Loukili H, Alami Younssi S, Albizane A, Bouhria M, Persin M, Larbot A. 2004c. Filtration of solution containing heavy metals and dyes by means of ultrafi ltration membranes deposited on support made of Moroccan clay. Desalination 168:301–306. Saffaj N, Alami Younssi S, Albizane A, Messaoudi A, Bouhria M, Persin M, Cretin M, Larbot A. 2003. Préparation de membranes d’ultrafi ltration de faible coût à base d’oxyde de titane et d’aluminate de zinc sur support cordiérite. Application à la fi ltration de solutions ioniques. Récents Progrès en Génie des Procédés 89:507–514. Saffaj N, Alami Younssi S, Persin M, Cretin M, Albizane A, Larbot A. 2005. Processing and characterization

of TiO2/ZnAl2O4 ultrafi ltration membranes deposited on tubular support prepared from Moroccan clay. Ceramics International 31:205–210. Van Gestel T, Vandecasteele C, Buekenhoudt A, Dotremont C, Luyten J, Leysen R, Van der Bruggen B, Maes G. 2002. Salt retention in nanofi ltration with

multiplayer ceramic TiO2 membranes. J. Membr. Sci 207:379–389.

320 WATER QUALITY RESEARCH JOURNAL OF CANADA

Philip H. Jones Award

A student award is given at each regional symposium for the best oral presentation. This award was created as a memorial to the late Philip H. Jones, who was a founding and longstanding member of the CAWQ. It includes a cash prize of $200 at the closing ceremony of the symposium; a one-year membership in the CAWQ, including a subscription to the Water Quality Research Journal of Canada; the publication of the name of the winning student in an issue of the journal and on the CAWQ website; a certifi cate acknowledging his/her performance; and a formal invitation from the CAWQ president to submit his/her work for peer review and eventual publication in the Water Quality Research Journal of Canada.

For the 24th Eastern Canadian Symposium on Water Quality Research held on November 7, 2008 in Montreal, Quebec, the Philip H. Jones award winner was:

Jason F. Carpenter Civil Engineering Department Université Laval Quebec City, Quebec for a presentation titled “Reducing pollutant discharge into urban rivers by real-time control of the stormwater retention time in a stormwater pond.”

For the 44th Central Canadian Symposium on Water Quality Research held on February 23 & 24, 2009 in Burlington, Ontario, there were two Philip H. Jones award winners: 1)

Tony Tsui Department of Civil Engineering University of Toronto Toronto, Ontario for a presentation titled “Statistical Analysis of Ottawa Lead in Tap Water Data from 1997 to 2008,” and 2)

Harris R. Switzman School of Geography and Earth Sciences McMaster University Hamilton, Ontario for a presentation titled “Quantifying Groundwater Discharge at the Lake-Groundwater- Beach Interface at Recreational Beaches.”

v WATER QUALITY RESEARCH JOURNAL OF CANADA

We thank the following individuals and their organizations for their role in reviewing the manuscripts submitted to Volume 43 of the Water Quality Research Journal of Canada. We also thank those reviewers whose names may have been omitted.

Abbassi, BE Al Balqa’ Applied University Allen, G University of Toronto Amon, J Wright State University Anderson, WB University of Waterloo Antoniades, D Pavillon Abitibi-Price Université Laval Baird, D University of New Brunswick, Environment Canada Barbeau, B Department of Civil, Geological and Mining Engineering, École Polytechnique de Montréal Barica, J United Nations University Basu, O Carleton University Blais, J-F Universite du Quebec Blume, T Technical University Hamburg-Harburg, Department of Sanitary and Environmental Engineering Blumenstein, M Griffi th University, Australia Bouchard, C Universite Laval Brar, S Institut national de la recherche scientifi que Brooks, T Health Canada Brown, J Carollo Engineers Brownlee, B AEMRD Chambers, J Chambers Institute of Environmental Science, Murdoch University Colussi, J California Institute of Technology Connelly, M British Columbia Institute of Technology Corkum, J Environment Canada Daigle, A INRS-ETE Darbi, A University of Regina Darby, J Department of Civil and Environmental Engineering, University of California Dolan, J CNRS Dubé, M University of Saskatchewan Emelko, M Environmental Engineering, University of Waterloo Ernst, B Environment Canada Fairchild, W Fisheries and Oceans Canada Falsanisi, D Technical University of Bari Fries, K CH2M Fu, G formerly of Okanagan University College Gagnon, G Centre for Water Resources Studies, Dalhousie University Galvez-Cloutier, R Université Laval Garrigues, P Universite de Bordeaux Gobeil, C INRS Godin, P Environment Canada Heinonen-Tanski, H Department of Environmental Science, University of Kuopio Hoffman, R Department of Civil Engineering, University of Toronto Holdway, D University of Ontario Institute of Technology Hunter, F Brock University Johnson, D Colorado State University

vi WATER QUALITY RESEARCH JOURNAL OF CANADA

Kok, S Environment Canada Landine, R University of New Brunswick Lemieux, F Health Canada Li, J Stantec Consulting Ltd. MacLatchy, D Wilfrid Laurier University Mahendraker, V ZENON Membrane Solutions Mahmood, T Paprican Mariñas, B Department of Civil Engineering, University of Illinois McFadyen, S. Health Canada McMartin, D University of Regina Mercier, G Institut National de la Recherche Scientifi que Monette, F Institut des sciences de l’environnement Montagnes, D University of Liverpool Muyibi, S International Islamic University Narasiah, S University of Sherbrooke Niquette, P Universite du Quebec Parrott, J Environment Canada Piggott, A Environment Canada Reebs, S Universite de Moncton Rousseau, A IRNS-ETE (Institut national de la recherche scientifi que - Centre Eau, Terre & Environnement) Roy, R Fisheries and Oceans Canada Santoro, D Trojan Technologies Seidou, O Universtiy of Ottawa Siddique, T University of Alberta Singh, K University of New Brunswick Solé, JD University of Girona, Spain Soupir, M Virginia Tech Springthorpe, S University of Ottawa Srinivasan, PT MWRD of Greater Chicago St Hilaire, A INRS-ETE, Université du Québec Straub, D McGillUniversity Surette, C Université de Moncton Swain, L Government of British Columbia Sweetman, J Parks Canada Templeton, M Imperial College London Thirunavukkarasu, O Saskatchewan Environment Turner, C University of Texas at El Paso (UTEP) Unc, A New Mexico State University van der Kraak, G University of Guelph Van Duin, B Westhoff Engineering Resources, Inc. van Sempvoort, D Environment Canada Viraraghavan, T University of Regina Watt, E Queen’s Univ. (Emeritus) and XCG Consultant Yan, G Alberta Capital Region Wastewater Commission Zhu, J University of Victoria

vii WATER QUALITY RESEARCH JOURNAL OF CANADA

KEY WORD INDEX FOR VOLUME 43 (2008)

Acidifi cation: 85 E. coli: 77, 129 Activated sludge: 201 NAR: 137 Air bubbles: 239 Eaux usées: 219 Alum: 231 Ecotoxicology: 257 Aluminium sulfate: 219 EDTA: 313 Analytic hierarchy process and uncertainty: 1 Effl uent: 161 Anoxic: 211 Endocrine Ash: 231 disrupting chemicals (EDCs): 173 Atrazine: 265 disruptor: 275 Endotoxins: 291 Bacteriophage: 69 Escherichia coli: 63 Bench-scale: 305 Estradiol, ethoxyresorufi n-O-deethylase (EROD): Bioassay: 275 275 Biofi lm: 249 Eutrophication: 85 Biofi ltration: 211 Extract: 231 Biogas: 211 Biopolymer: 219 Factor analysis: 111 Biopolymère: 219 Factors for Trihalomethanes formation: 189 Biotic: 69 Faecal coliforms: 283 Biotracer: 137 Fathead minnow: 257 Biotrickling: 211 Fish: 161, 265, 283 Bisphenol-A: 265 Flocculation: 231 Bleu de méthylène: 313 Flocs: 239 Floculant: 219 Cations: 201 Food vacuoles: 69 Cavitation: 23 Frequency: 183 Chitosan: 219 Fuzzy synthetic evaluation: 1 Chitosane: 219 Chlorine: 23, 63, 291 Groundwater: 111 Chlorine dioxide: 11, 63 Chloropicrin: 11 Haloacetaldehydes: 11 Chrome hexavalent: 313 Haloketones: 11 Ciliate: 69 Human health: 283 Climate change: 85, 145 Hydraulics: 137 Coagulant: 219, 231 Hydrocarbons: 99 Constructed wetlands: 137 Hydrogen sulphide: 211 Contaminated soils: 99 Cr(VI): 313 Inactivation: 69 Cytochalasin B: 69 Industrial and mining pollution: 99 Industrial wastewater treatment: 173 Defl occulation: 201 Inorganic DBPs: 11 Denitrifi cation: 211 Intermittent supply: 249 Diatoms: 85 Irrigation water: 145 Digestion: 305 Disinfection: 23, 55, 77 Kraft: 161 approach: 1 Dissolved Air Flotation (DAF): 239 Life cycle bioassasy: 257 Dissolved oxygen transient: 201 Lithium tracer: 137 Distribution system: 55 Drinking water: 11, 55, 77, 249 Manure type: 129

viii WATER QUALITY RESEARCH JOURNAL OF CANADA

KEY WORD INDEX FOR VOLUME 43 (2008)

Membrane: 313 Simultaneous variation: 189 Methoxychlor: 265 Sitosterol, E-: 173 Methylene blue: 313 Size assortative reproduction: 257 Microscope-particle counter: 239 Sludge: 305 Mixed function oxygenase (MFO): 275 Small systems: 77 Modelling microbial inactivation: 47 Soil erosion: 121 Monochloramine: 63 Sonication: 37 MS2: 69 Sonochemical reactor: 183 Multicriteria decision making: 1 Sphingomonads: 249 Multivariate statistical analysis: 189 Stormwater Municipal effl uent: 37 pond: 145 wastewater: 275 reuse: 145 Sulfate d’aluminium: 219 Nitrate runoff: 121 Sustainable agriculture: 121 Northwestern Mexico: 111 Synergy: 63 Nova Scotia: 85 Nutrient(s): 111, 305 Testosterone: 275 Tetrahymena: 69 Organic DBPs: 11 Thermomechanical: 161 Ozone: 291 Thickening: 305 Tissue distribution: 265 Paleolimnology: 85 Trace metals: 111 Particle associated bacteria: 47 Treatment: 161 Particles: 37, 55 time: 183 Pathogens: 77 Trickling fi lter: 37 PCR: 249 Turbidity: 231 Peracetic acid: 47 Plant sterols/phytosterols analysis: 173 Ultrafi ltration: 313 Plantain: 231 Ultrasound: 23, 37 Polymer: 305 Ultraviolet: 23 Polysaccharides: 201 Ultraviolet light: 63 Potato rotation: 121 UV: 37, 291 Power: 183 Primary: 37 Virus: 69 Principal component analysis: 189 Proteins: 201 Wastewater: 23, 219 Pulp and paper mill effl uents: 173 Wastewater disinfection: 47 Pulsed arc electrohydraulic discharge: 77 Water quality: 121, 145, 283 Whole-body autoradiography: 265 Rainbow trout: 275 Rainfall simulation: 129 Yellow colonies: 249 Reactive yellow dye: 183 Reproduction: 161 Riverbank: 99 Runoff: 129

Sewage: 283 Sewage effl uent: 275 Sex steroids: 275 Silylation: 173

ix WATER QUALITY RESEARCH JOURNAL OF CANADA

AUTHOR INDEX FOR VOLUME 43 (2008)

Acosta, B. 111 Gibson, J.H. 23 Parrott, J.L. 275 Albizane, A. 313 Ginn, B.K. 85 Pinheiro, M.D.O. 69 Allen, D.G. 201 Gorczyca, B. 239 Pollock, M.S. 257 Alliu, Y.D. 231 Gordon, R. 137 Pollock, R.J. 257 Alvarez-Castañeda, S.T. 111 Power, M.E. 69 Aranda-Rodriguez, R. 11 Hähni, M. 99 Arnold, A.J. 77 Hall, E.R. 173 Rand, J.L. 63 Azam, K. 183 Hashwa, F. 249 Rouleau, C. 265 He, J. 145 Babineau, D. 219 Hofmann, R. 55 Saint-Laurent, D. 99 Basu, O.D. 305 Santillan, C.A. 77 Baxter, C. 55 Jamieson, R. 137 Seto, P. 23, 211 Beasley, B.W. 129 Jay, B. 11 Shupe, G. 63 Béland, M. 211 Slawson, R. 69 Benoit, F.M. 11 Kidd, K.A. 283 Smol, J.P. 85 Bérubé, V. 161 Klassen, P. 239 Soreanu, G. 211 Bols, N.C. 69 Kohli, J. 265 Squires, A.J. 257 Boutilier, L. 137 Koudjonou, B. 11 St-Laurent, J. 99 Brimacombe, M. 121 Kovacs, T. 161 Brito-Castillo, L. 111 Tahiri, S. 313 Burney, J.R. 121 Lake, C. 137 Thienpont, J.R. 85 Butler, B.J. 69 LeBel, G.L. 11 Tokajian, S. 249 Leduc, R. 219 Trowbridge, D. 275 Cairns, W. 37 Lee, L.E.J. 69 Chahid, E.G. 313 Lee, L.H. 77 Uribe, S.P. 291 Champagne, P. 1, 189 Liberti, L. 47 Chartray, D. 219 Loomer, H. 283 Valeo, C. 145 Chivers, D.P. 257 Loukili, H. 313 Verma, S. 275 Chowdhury, S. 1, 189 Lynn, D.H. 69 Vickers, T. 283 Chu, A. 145 Cumming, B.F. 85 MacDonald, A.J. 305 Wojcicka, L. 55 MacRae, A.H. 121 Wurlz, J. 111 Da Silva Baptista, I.F. 291 Mahmood-Khan, Z. 173 Dayeh, V.R. 69 Mahvi, A.H. 183 Yong, H.N. 23, 37 Dehghani, M.H. 183 Majouli, A. 313 Younssi, S.A. 313 Dubé, M.G. 257 Mao, T. 37 Martel, P. 161 Zhang, Y. 201 Edmonson, K. 211 Mazer, B. 291 Edwards, L. 121 McAslan, A. 283 Emelko, M.B. 77 McLellan, P.J. 189 McMaster, M.E. 275 Falletta, P. 211 Méndez, L. 111 Falsanisi, D. 47 Mesdaghinia, A.R. 183 Farah, M. 249 Miller, J.J. 129 Farnood, R.R. 23, 37 Fisher, S.E. 257 Nasseri, S. 183 Neumann, N.F. 145 Gagnon, G.A. 63 Notarnicola, M. 47 Gardner, S.C. 111 Gehr, R. 47, 291 Oladoja, N.A. 231

x WATER QUALITY RESEARCH JOURNAL OF CANADA

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