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Polycyclic musk compounds have been determined in the environmental matrices and in biological tissues in the last decade yet no reference to their chronic toxicity in the marine environment. In the present research, the clams, Ruditapes philippinarium were exposed to 0.005, 0.05, 0.5, 5 and 50 µg/L of galaxolide (HHCB) and tonalide (AHTN) for 21 days. A battery of biomarkers related with xenobiotics biotransformation (EROD and GST), oxidative stress (GPx, GR and LPO) and genotoxicity (DNA damage) were measured in the digestive gland tissues. HHCB and AHTN significantly induced EROD and GST enzymatic activities (p<0.05) at environmental concentrations. These substances induced GPx activity and significantly inhibit GR activity (p<0.05). All concentrations induced significant increase of LPO on day 21 for both substances leading to DNA damage. Although these substances have been reported none acutely toxic, this study have shown they induce oxidative stress and consequently, genetic strands break in marine organisms.
Fragrances are low molecular weight (< 300 g/mol) chemical compounds which are applied in personal care products such as body lotions, washing soap and detergents, perfumes, toothpastes, various cosmetics, etc (Reiner and Kannan, 2011). They are derived from musks – a range of natural and synthetic substances. Natural musks were extracted from fatty tissues of plants and animals. However, due to increased demand and the consequential high cost coupled with uncertainty in supply of natural musk, synthetic musk compounds (SMC) were formulated as a replacement (Rimkus, 1999). The nitro derivatives which are composed of methylated nitrates and acetylated benzene rings were the first to be synthesized and used on a commercial scale (Sommer, 2004). They were however banned due to their toxicity and persistence in the environment. The polycyclic musk compounds (PMC), unlike the nitro counterparts, are bi-cyclic aromatic structure which comprised of acetylated and highly methylated pyran, tetralin and indane skeletons (Sumner et al., 2010). Although the industrial synthesis of this group are moderately difficult, they are however, considered very important fragrance materials in perfumery owned to their ability to bind to fabrics and their characteristic musky scent (Swidish Society for Nature Conservation, 2000). Of this category, galaxolide (HHCB) and tonalide (AHTN) are the most consumed PMC accounting for 95% of total market volume of fragrance materials (Balk and Ford, 1999; Pedersen et al., 2009). The estimated quantity of these substances used in 1996 was approximately 8000 tone (Chen et al., 2010). In the United States, all fragrances consumed has been doubled since 1990 (Roosens et al., 2007) and increased by 25% between 1996 and 2000, from estimated quantity of 5200 to 6500 tone (Peck et al., 2006).
SMC were first determined in the environmental matrices in Tama river, near Tokyo Japan in 1981 (Peck et al., 2006). These compounds have been found in various environmental matrices: air, water and sediment, at relevant environmental concentrations (Fromme et al., 2001; Peck et al., 2006; Peck and Hornbuckle, 2006). Owing to their external applications, the main route of entrance is through sewage treatment plants and by atmospheric deposition (Fromme et al., 1999; Ramskov et al., 2009). Most sewage treatment plants are not adapted to eliminating completely SMC from municipal and industrial waste water. Investigations have shown that only about 50 to 90 percent of the total SMC are eliminated from sewage treatment plants while the rest enter the receiving rivers and oceans via sewage outfall (Heberer, 2002; Lee et al., 2010) and are diluted along the river gradient downstream (Ricking et al., 2003). This account for the higher concentrations measured at sewage treatment effluents. In Germany for example, HHCB and AHTN concentrations up to 13330 and 4360 ng/L respectively measured at sewage effluent were highest in surface water downstream at 1590 ng/L and 530 ng/L respectively (Fromme et al., 2001). Similarly, Sumner et al. (2010) measured up to 2089 and 530 ng/L of HHCB and AHTN at sewage effluent, and 30 and 15 ng/L downstream in Tamar Estuarine respectively. The mean concentrations of HHCB and AHTN measured at three sewage effluents off the Bay of Cadiz in south of Spain ranged from 5603 ng/L to 592 ng/L, representing more than the 50% of the contaminants measured in the effluents, which discharge to marine environments (Díaz-Garduño et al., 2017).
There are few reports on measured environmental concentrations of SMC in marine environment (Bester et al., 1998; Sumner et al., 2010). This is very disturbing for two reasons: firstly, because of the increasing consumption of these compounds globally; and secondly, many cities are located on the coast and thus have sewage effluent arriving over a short distant into the marine environment. However, reports have shown bioaccumulation of these compounds in marine organisms (Kannan et al., 2005; Moon et al., 2011, 2012). Moon et al. (2012) studied the concentration and accumulation profiles of polycyclic aromatic hydrocarbon and SMC in liver tissues and bubbler from minke whales and common dolphin from Korean coastal waters. They reported that in all the samples of liver tissues and bubbler, HHCB were found to be predominant with no exception. The concentrations in ng/g lipid weight found in the liver tissues and bubbler of both organisms ranged from <2.3 to 169 and <2.3 to 50; and 24 to 187 and 19 to 72 of HHCB and AHTN respectively. The highest concentrations of both compounds were found in the liver and bubbler of minke whales. Most of the studies focused on bioconcentration but to understand the biomagnification potential of these substances and to identify those species that accumulate higher concentrations through their diet and habitat, Nakata et al., (2007) measured synthetic musks in lugworm, mussel, crustacean, fish, marine birds and mammals from tidal flat and shallow water areas of Ariaka Sea, Japan. In all the samples analysed, HHCB and AHTN were the dominant compounds while the nitro counterparts were undetected. An earlier investigation of oysters in 61 locations around Japanese coastal waters reported the predominance of HHCB and AHTN (Breitholtz et al., 2003). HHCB concentrations in organisms from the Ariaka Sea were 3 to 10 folds higher than those of AHTN and the highest concentrations were found in clams, ranging from 258 ng/g (lipid wt.) to 2730 ng/g (lipid wt.), followed by crustaceans and fish in the tidal flat.
The toxicity of PMC, especially HHCB and AHTN has been performed with species from different trophic levels and life stages to ascertain their toxicity. The effects of HHCB and AHTN on larvae development, juvenile growth and survival of different organisms have been recorded (Wollenberger et al., 2003; Breitholtz et al., 2003; Carlsson and Norrgren, 2004; Gooding et al., 2006; Pedersen et al., 2009). Although many SMC are not regarded as acutely toxic to organisms at environmental relevant concentrations, sub-lethal effects of ecological relevance have been detected (Breitholtz et al., 2003; Gooding et al., 2006; Pedersen et al., 2009; Chen et al., 2011). Most of these studies are performed for freshwater environment and are for short term exposure.
Materials and Methods
2.1. Selection of Polycyclic Musks
HHCB and AHTN were purchased from Sigma Aldrich, Spain. The characteristics of these substances are presented in table 1. The concentrations used were carefully selected based on reported concentrations measured in different environmental matrices.
2.2. Acclimation and maintenance conditions
The species, Ruditapes philippinarum was selected for this study. A total of 360 specimens were purchased from an aquaculture farm in the bay of Cadiz, Spain. The average size of the specimens was 42±0.9mm. They were transported immediately to the laboratory for acclimation and kept in a 300L aquarium for seven days. The aquarium was supplied with constant aeration and the specimens were feed ad libitum once every day. The physical and chemical parameters in the aquarium were monitored and controlled in a photoperiod of 12 h light/12 h dark; temperature 15±1 0C; salinity 34.6±0.3%0; pH 7.8 – 8.2; dissolved oxygen >5mgL-1.
2.3. Experimental Approach
Concentrations of HHCB and AHTN were based on environmental measurement (0.005, 0.05, 0.5, 5.0 and 50.0 μgL-1) for 21 days exposure in a semi-static renovation bioassay. The musks were dissolved in dimethyl sulfoxide (DMSO) in glass vials. These stock solutions were stored in dark bottles at 4oC in the refrigerator. They were then diluted with nanopore water each time to obtain the desired concentrations. The bioassay was conducted in 10 L capacity rectangular glass aquarium in duplicates including sea water control and solvent control (DMSO) to ensure there was no solvent effect (Aguirre-martínez et al., 2016). The glass aquariums were filled with 8 L of sea water spiked with the contaminants. The bioassay was renewed every 3 days with the water siphoned out of the aquariums and properly cleaned and refilled. Volumes of freshly prepared solutions were added at each renewal period to expose the organisms to desired concentrations. Physiochemical parameters were akin to the acclimation condition reported above. Samples of clams were collected on days 3, 7, 14 and 21 and immediately stored in the refrigerator at -80 0C in the laboratory.
2.4. Water Samples Collection and Analyses
Water samples used in bioassays experiments were collected on days 0 and 3, using amber bottles and immediately stored at -20 0C prior to analyses. The target compounds (HHCB and AHTN) were measured and quantified to analyse the water samples spiked with HHCB and AHTN on days 0 and 3. The methodology used was the stir bar sorptive extraction (SBSE) following a modification of methodology described by Pintado-Herrera et al. (2014). Prior to use, all polydimethylsiloxane bars (PDMS, 10 mm x 0.5 mm) were preconditioned by soaking them in a mixture of acetonitrile/methanol (80:20, v/v). Later, these bars were placed in amber glass flasks containing the aqueous samples (500 ml), 1 µg l-1 of benzophenone d10 was also added to determine possible fluctuations during the extraction and analysis procedures and stirred at 900 rpm during 4 h at room temperature. After extraction, the bars were desorbed by liquid desorption (LD); bars were sonicated during 30 min in vials containing 500 µL of ethyl acetate. Then, gas chromatography (SCION 456-GC, Bruker) and mass spectrometry (SCION TQ from Bruker with CP 8400 Autosampler) were used to identify and quantify the compounds. Capillary gas chromatography analysis was carried out on a HP-5MS column (30 m×0.25 mm i.d.×0.25 μm film thickness of 5 % phenyl, 95 % polydimethylsiloxane), keeping the helium carrier gas flow at 1 mL min−1. The mass detector acquired in multiple reaction monitoring (MRM) mode. Details of the detection methodology can be found in Pintado-Herrera et al. (2016). Calibration curves were constructed for each compound in the range of 0.01–20 μg L-1. Method limits of quantification (mLoQ), calculated using a signal-to-noise ratio 10 to 1, respectively, for water samples were lower than 0.04 ng L−1. The recovery of the method was higher than 85% for both analytes.
2.5. Samples Preparation
Samples stored in refrigerator were defrosted in ice and digestive gland tissues were extracted. Digestive glands of 3 clams from each aquarium were pooled together and stored at -80oC prior to homogenization. Homogenization buffer was prepared with 100mM NaCl, 25mM HEPES salt, 0.1mM EDTA and 0.1mM DTT. Pooled samples were homogenized and samples for biochemical analyses were centrifuged to obtain supernatant portions at speed of 15.000 x g for 20 minutes at 4oC (S15) and 3.000 x g for 20 minutes at 4oC (S3). (Bradford, 1976) methodology was adapted to determine corresponding total protein (TP) and values expressed as mg/ml for different extracts (HF, S3 and S15). All biomarkers were measured using a kinetic microplate reader, Infinite® M200.
2.6. Biomarker Analyses
2.6.1 Ethoxyresorufin O-deethylase (EROD) activity
Mixed function oxidase activity was measured using EROD assay adapted from fingerlings of rainbow trout (Gagné and Blaise, 1993) to clams in the present work. 50 μl of the S15 sample (25 μl sample +25 µl of MilliQ) was coated in dark 96 flat bottom wells microplate and 160 μl 7-ethoxyresorufin, 10 µl of reduced NADPH, in 100 mM K2HPO4 buffer at pH 7.4 was added. The reaction was sparked up by the addition of NADPH and continued for 60 mins (15 mins intervals) at 30oC. 7- hydroxyresorufin was determined fluorometrically using 516 nm excitation and 600 nm emission filters. Calibration was then achieved through a standard calibration curve developed with concentrations of resorufin. Results were normalized to total protein (TP) content. Results were expressed as pmol/min/mgTP.
2.6.2 Glutathione-S-Transferase (GST) activity
Determination of GST activity was adopted from Boryslawskyj et al. (1998) procedure. In brief, 15 µl of S15 samples were added to 200 µl solution of 1mM 1-chloro-2.4-dinitrobenzene, 10mM HEPES salt, 125mM NaCl and 1mM glutathione reduce (GSH) normalized at pH 6.5. The mixtures were placed in a transparent microplate containing 96 flat bottom wells. Absorbance based on the appearance of the glutathione conjugate was measured at 340 nm at 0, 5, 10, 15, 20, 25 and 30 minutes. The results were expressed as optical density per minute per milligram of total protein (μg/min/mg/TP).
2.6.3 Glutathione peroxidase (GPx) activity
Glutathione peroxidase activity was measured following oxidation of NADPH to NADP. The procedure applied was adapted from Mcfarland et al. (1999). In a transparent 96 flat bottom wells microplate, 10 μl of the S15 sample was diluted with 10 μl of MilliQ and measured spectrophotometrically at 340 nm, at 3 s intervals, for 3 mins using 1 mM cumene hydrogen peroxide as substrate. The decrease in NADPH absorbance was indicative of GPx activity. Results were expressed as nmol/min/mgTP.
2.6.4 Glutathione Reductase (GR) activity
The method adapted by Martin-Diaz et al. (2007) was used in measuring the activities of GR. 10 µl of the S15 samples and 10 µl of MilliQ were placed in a transparent microplate containing 96 flat bottom wells. 200 µl of the incubated daily assay was added (10mM oxidized glutathione and 1mM NADPH in 200 mM sodium phosphate buffer with pH of 7.6) to the samples. The consumption of NADPH produces a decrease in absorbance, which is directly proportional to the glutathione reductase activity in the samples. The absorbance was measured spectrophotometically at 340 nm, every 2 minutes for 10 minutes at 30oC.
2.6.5 Lipid Peroxidation (LPO)
LPO was measured following the thiobarbituric acid reactive substances (TBARS) method developed by Wills (1987). Oxidative stress results in malondialdehyde (MDA) production from the degradation of initial products of free radical attacks on fatty acids (Janero, 1990). MDA reacts with 2-thiobarbituric acid producing tetramethoxypropane (TMP). This can be measured spectrophotometrically allowing the indirect measurement of MDA. In a 1.5 mL eppendorf, standard solutions and homogenate samples were prepared separately. To set a standard curve of TMP, standard solution of TMP 0.001% was prepared and a serial dilution of the solution with distil water (at 0 + 150; 6 + 144; 15 + 135; 30 + 120; 40 + 110; 60 + 90; and 100 + 50) µl respectively were placed in eppendorfs. 300µl of trichloroacetic acid (TCA) 10%, 1 mM FeSO4, and 150 µl of thiobarbituric acid (TBA) 0.67% were added to diluted homogenate (90µl of the sample + 60µl of MilliQ Water). Standard and homogenates were incubated at 70oC for 10 minutes in P Selecta Heater®. Thereafter, 200 µl of the standard solution for the TMP standard curve and 200 µl of the samples precipitate were pipetted into dark microplate containing 96 flat bottom wells. Fluorescence was measured at 516 (excitation) and 600 (emission) filters and results were expressed as µgTBARS/mgTP.
2.6.6 DNA damage
DNA precipitation assay methodology is based on 2% SDS-KCl precipitation of DNA-protein crosslink, which uses fluorescence to quantify the DNA strands (Olive, 1988; Gagne et al., 1995). When DNA breaks because of exposure to toxic chemicals, the strands are released from cellular protein into the supernatant when centrifuged at low speed (Olive, 1988). It becomes possible to quantify the amount of double and single stranded DNA at the end of the assay (Aguirre-Martinez et al., 2016). A volume of 25 µl of homogenate was mixed by inversion with 200 µl of SDS 2% prepared with 10mM EDTA, Tris-Base and 40 mM NaCl. 200 µl of 0.12 KCl was added and mixed by inversion. The mixture was incubated for 10 minutes at 60oC, cooled at 4oC for 30 minutes and centrifuged 8000 x g at 4oC for 5 minutes. For DNA calibration, 1 mg Salmon Sperm genomic DNA was dissolved in 1 ml TEIX (Tris-HCl and EDTA at pH 8.0) as standard. In a 96 flat bottom wells dark microplate, 50 µl of the supernatant was added to 150 µl of Hoescht dye 0.1 µg mL-1 diluted with sodium cholate containing 0.4M NaCl, 4mM sodium cholate and 0.1M Tris-Acetate (pH 8.5). Fluorescence was measured at 360 nm (excitation) and 450 nm (emission) filters against blank containing similar constituents, without homogenate. Results were expressed as µgDNA/mgTP.
2.7. Statistical Analysis
Data from biomarker responses were analyzed using the SPSS/PC + statistical package®. Normality of the data and homogeneity of variance were analyzed prior use of parametric test. Significant differences between controls and organisms exposed to polycyclic musks treatments were determined using one-way analysis of variance (ANOVA); the data were not transformed and Dunnette’s multiple comparison tests were performed. Spearman’s rank order of correlation test was used to obtain pair wise correlations and significant levels were set at p < 0.05 and p < 0.01.
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