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Effects of BDE-99 on hormone homeostasis and biochemical parameters in adult male rats
Food and Chemical Toxicology 48 (2010) 2206–2211
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Effects of BDE-99 on hormone homeostasis and biochemical parameters in adult male rats Virginia Alonso a,b, Victoria Linares a,b, Montserrat Bellés a,b, María Luisa Albina a,b, Andreu Pujol b, José L. Domingo a,*, Domènec J. Sánchez a,b a b
Laboratory of Toxicology and Environmental Health, School of Medicine, IISPV, Universitat Rovira i Virgili, Sant Llorens 21, 43201 Reus, Catalonia, Spain Physiology Unit, School of Medicine, IISPV, Universitat Rovira i Virgili, Sant Llorens 21, 43201 Reus, Catalonia, Spain
a r t i c l e
i n f o
Article history: Received 12 March 2010 Accepted 13 May 2010
Keywords: BDE-99 Adult rats Oxidative stress Endocrine disruption Hepatotoxicity Nephrotoxicity
a b s t r a c t In this study, we evaluated the effects of BDE-99 on hormone homeostasis, as well as in urinary and serum biochemical parameters of adult male rats. Animals (10 per group) received BDE-99 by gavage at single doses of 0, 0.6 and 1.2 mg/kg. Forty-five days after BDE-99 exposure, urine and serum samples were collected for hormonal and biochemical analysis. Oxidative stress (OS) markers in erythrocytes, plasma and urine were also evaluated. Urinary excretion of total protein significantly increased following BDE-99 exposure, while lactate dehydrogenase (LDH), c-glutamil transferase (GGT), and N-acetylglucosaminidase (NAG) activities significantly decreased. Liver toxicity was evidenced by elevated serum activities of the enzymes glutamic oxaloacetic transaminase (GOT), glutamic pyruvic transaminase (GPT) and alkaline phosphatase (ALP). Following BDE-99 administration, OS markers in erythrocytes showed an increase in superoxide dismutase (SOD) activity, and a reduction in glutathione reductase (GR) activity. In urine, isoprostane levels increased after BDE-99 exposure. The hormonal analysis showed a significant decrease in testosterone and progesterone levels. These results support the hypothesis that BDE-99 interacts with hormonal response. Moreover, BDE-99 administration to adult male rats showed signs of renal and hepatic toxicity. Ó 2010 Elsevier Ltd. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs), a group of brominated flame retardants formed by 209 congeners, are used as additives in plastics, textiles, foams and electronic appliances. These compounds have a high lipophilicity, are environmental persistent, and possess a notable potential of bioaccumulation. PBDEs are found in air, sediments, biota and human tissues (Darnerud et al., 2001; Law et al., 2006; Sjödin et al., 2008). For the general popula-
Abbreviations: ALP, alkaline phosphatase; ALT, alanine aminotransferase; PCBs, polychlorinated biphenyls; AST, aspartate aminotransferase; BDE-99, 2,20 ,4,40 ,5pentabromodiphenyl ether; CAT, catalase; ChE, cholinesterase; ER, estrogenic receptors; GGT, c-glutamil-transferase; GOT, glutamic oxaloacetic transaminase; GPT, glutamic pyruvic transaminase; GPx, glutathione peroxidase; GR, glutathione reductase; GSH, reduced glutathione; GSSG, oxidized glutathione; LDH, lactate dehydrogenase; NAG, N-acetylglucosaminidase; OS, oxidative stress; PBDEs, polybrominated diphenyl ethers; POCs, persistent organochlorinated compounds; ROS, reactive oxygen species; SOD, superoxide dismutase; T3, triiodothyronine hormone; T4, thyroxine hormone; TBARS, thiobarbituric acid reactive substances; TSH, thyroid-stimulating hormone. * Corresponding author. Tel.: +34 97759380; fax: +34 977 759322. E-mail address: [email protected] (J.L. Domingo). 0278-6915/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.fct.2010.05.048
tion, the main routes of exposure to PBDEs, particularly the lower brominated congeners, is through the diet, as it also occurs with dioxins and PCBs. Consequently, consumption of foodstuffs such as fatty fish from contaminated sources is a major way of exposure to PBDEs (Bocio et al., 2003; Domingo, 2004; Darnerud et al., 2006; Domingo et al., 2006, 2008; Martí-Cid et al., 2007; Perelló et al., 2009). In previous studies on the levels of PBDEs in a number of foodstuffs, we analyzed individually the following six tetrathrough heptabrominated congeners: 47, 99, 100, 153, 154 and 183 (Domingo et al., 2006, 2008; Martí-Cid et al., 2007). For most food groups, congeners BDE-47 and BDE-99 showed the highest concentrations. These specific congeners were analyzed taking into account that they are the most abundant in the environment, being also the easiest of measuring (Domingo, 2004). In addition, these congeners, together with BDE-28 (triBDE) and BDE-209 (decaBDE), are those recommended by the European Food Safety Agency in monitoring programs for feed and food (EFSA, 2006). From a toxicological point of view and although limited data are currently available on these compounds, new toxicology data can be expected in the near future (EFSA, 2006). Human tissue concentrations of persistent organochlorinated compounds (POCs), such as polychlorinated biphenyls (PCBs) have
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211
decreased since their restriction in the 1970s (Meironyte and Noren 2001; Sjödin et al., 2004). In contrast, monitoring studies suggest that human tissue levels of PBDEs have increased in recent years (Lind et al., 2003; Sjödin et al., 2004; Johnson-Restrepo et al., 2005; Wang et al., 2008). Although there is strong evidence that PCBs and other POCs may cause adverse health effects in mammals (WHO, 2002), the potential toxicity of PBDEs in humans remains still to be thoroughly ascertained (Darnerud et al., 2001; Darnerud, 2003; Costa et al., 2008). Structurally, PBDEs are similar to PCBs, a different toxicity for the 209 possible PBDE congeners can be also expected. Among PBDE congeners, 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE-99) is one of the most persistent, being also detected in most environmental and food samples (Domingo, 2004; Domingo et al., 2006, 2008). To date, only data on a few experimental in vivo studies on BDE99 toxicity have been reported, while mechanistic information is also clearly lacking. BDE-99 has been found to elicit various adverse effects including endocrine disruption (Hakk et al., 2002), neurobehavioral toxicity (Eriksson et al., 2002; Cheng et al., 2009), and/or developmental and reproductive toxicity (Kuriyama et al., 2005; Lilienthal et al., 2006). However, most experimental studies have been focused on pre- and perinatal exposure, lacking information in adult animals. Because PBDEs have structural similarities to the hormones thyroxine (T4) and triiodothyronine (T3) (Hamers et al., 2006), concern has been raised regarding their effects on thyroid function. Although the mechanism of endocrine disruption of PBDEs has not been established yet, recently Turyk et al. (2008) found an association between PBDE exposure and increased T4 in adult male fish consumers. Moreover, little is known about PBDEs action on androgenic systems (Stoker et al., 2005). Studies on BDE-99 distribution in rats concluded that the major target for BDE-99 is adipose tissue followed by the liver where it is detoxified (Hakk et al., 2002; Chen et al., 2006). Some authors have reported an increase of hepatic microsomal enzymes activities in rats exposed to the penta-PBDE mixture DE-71 (Zhou et al., 2001; Szabo et al., 2009), while recent studies in human hepatocytes have suggested that the human liver will likely metabolize some BDE congeners (e.g., BDE-99) in vivo (Stapleton et al., 2009). On the other hand, toxic effects of BDE-99 on biochemical markers of renal function are not well known. Most studies in mammals are only based on observation of increased organ weight (ATSDR, 2004). Recently, we conducted in adult rats an investigation about renal effects of BDE-99 exposure. Results showed that BDE-99 exposure induced oxidative stress (OS) in kidneys with morphological alterations (Albina et al., 2010). In the present investigation, the role of OS as a mechanism for BDE-99 toxicity has been assessed. This hypothesis was based on the results of previous studies reporting that exposure to some persistent organic pollutants such as PCBs, could produce impairment through oxidative damage (Lehmann et al., 2007; Glauert et al., 2008). Although ROS serve as key signal molecules in physiological processes, they also play a role in pathological processes involving hormonal imbalances, neurodegenerative and reproductive impairments (Agarwal et al., 2003; Venkataraman et al., 2007). Erythrocytes can be regarded as circulating antioxidant carriers, reflecting exposure to ROS. They have been used as a model for the investigation of free-radical induced OS, because: (a) they are exposed to high oxygen tensions, (b) they are unable to replace damaged components, (c) the membrane lipids are composed partly of polyunsaturated fatty acid side chains, which are vulnerable to peroxidation, and (d) they have antioxidant enzyme systems (Kiruthiga et al., 2007). Taking the above into account, the purpose of the present study was to investigate the effects of BDE-99 on hormonal function of adult rats, as well as in blood and urine biomarkers.
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2. Materials and methods 2.1. Animals and chemical Adult male Sprague-Dawley rats (220–240 g) were obtained from Charles River (Barcelona, Spain). Animals were housed in plastic cages in a climate-controlled facility with a constant day–night cycle (light: 08:00–20:00 h) at a temperature of 22 ± 2 °C, and a relative humidity of 50 ± 10%. Food (Panlab rodent chow, Panlab, Barcelona) and tap water were offered ad libitum throughout the study. The use of animals and the experimental protocol were approved by the Animal Care and Use Committee of the ‘‘Rovira i Virgili” University (Tarragona, Catalonia, Spain). BDE-99 (98% pure) was purchased from LGC Standards S.L.U. (Barcelona). 2.2. Treatment After a quarantine period of 14 days, groups of 10 animals were given by gavage single doses of BDE-99 at 0.6 or 1.2 mg/kg of body weight (BDE 0.6 and BDE 1.2 groups, respectively). Rats in the control group received the vehicle only (corn oil). The choice of the BDE-99 doses was based on results of previous studies by Kuriyama et al. (2005). Forty-five days after BDE-99 exposure, rats were individually housed in plastic metabolism cages. Urines (24-h) were collected from animals of each group to determine biochemical parameters. After urine collection, rats were anesthetized by an intraperitoneal injection of ketamine–xylazine. Blood samples were collected from the vena cava to determine biochemical serum parameters and hormonal levels. Finally, rats were euthanized with an overdose of ketamine– xylazine. 2.3. Biochemical analysis Serum samples were analyzed for creatinine, urea, total protein, and uric acid concentrations, as well as for lactate dehydrogenase (LDH), cholinesterase (ChE), glutamic oxaloacetic transaminase (GOT), glutamic pyruvic transaminase (GPT) and alkaline phosphatase (ALP) activities using a Cobas Mira automatic analyzer (Roche Pharmaceuticals, Basel, Switzerland). In urine, volume, levels of creatinine, urea, total protein and uric acid, as well as LDH, c-glutamil-transferase (GGT), and N-acetylglucosaminidase (NAG) activities were also analyzed (Bellés et al., 2007). Urinary isoprostane levels (8-isoPGF2a) were determined as an indicator of lipid peroxidation (Cayman Chemical) (Shimizu et al., 2007). In erythrocytes and plasma, reduced (GSH) and oxidized (GSSG) glutathione were measured by the Hissin and Hilf method (1976), while thiobarbituric acid reactive substances (TBARS) were determined according to the Buege and Aust method (1978) modified for fluorimetric detection (Richard et al., 1992). In erythrocytes, superoxide dismutase (SOD), catalase (CAT) and glutathione reductase (GR) activities were determined according to previously described methods (Bellés et al., 2007), while glutathione peroxidase (GPx) activity was measured by the method of Wheeler et al. (1990). 2.4. Hormone analysis Serum samples were analyzed for thyroid-stimulating hormone (TSH), thyroxine (T4), triiodothyronine (T3), testosterone and progesterone concentrations using an Advia 2400 analyzer (Sakai et al., 2009). 2.5. Statistical analysis To evaluate the homogeneity of variances, the Levene test was used. When the variances of different treatment groups were homogeneous, ANOVA, and subsequently Tukey paired comparisons as a post hoc test, were applied to establish the level of significance among groups. If the variances were not homogeneous, the Kruskal–Wallis test and the Mann–Whitney U-test were used. The level of statistical significance for all tests was set at P < 0.05. All data were analyzed by means of the statistical package SPSS 15.0 (SPSS Sciences, Chicago, IL, USA).
3. Results The effects of exposure of adult male rats to BDE-99 on various urinary parameters are summarized in Table 1. Urinary excretion of total protein was significantly increased at 1.2 mg BDE-99/kg. LDH activity significantly decreased in the BDE-99 exposed groups, while GGT and NAG activities were significantly reduced at 1.2 mg BDE-99/kg. Although uric acid levels were increased by BDE-99, the differences with the control group were not statistically significant. Furthermore, there was a tendency to increase isoprostane levels in BDE-99 exposed groups. For the rest of urinary parameters no significant differences were noted.
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Table 1 Urinary parameters in adult rats treated with BDE-99.
Table 2 Serum parameters in adult rats treated with BDE-99.
Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals Urinary volume (ml/24 h)B Creatinine (mg/24 h)B Urea (mg/24 h)B Total protein (mg/24 h)A Uric acid (mg/24 h)B LDH (U/24 h/kg)B GGT (U/24 h/kg) B NAG (U/mg creatinine)B Creatinine clearance (ml/min)A Isoprostanes (ng 8isoPGF2a)B
10 11.75 ± 2.55
10 16.13 ± 7.71
10 13.86 ± 4.10
10 0.48 ± 0.07
10 0.48 ± 0.07
10 0.50 ± 0.11
16.44 ± 1.43
15.45 ± 1.81
15.22 ± 2.48
481.25 ± 143.37 29.22 ± 9.19a
398.75 ± 161.79 44.35 ± 23.58a
541.67 ± 189.41 126.30 ± 49.37b
32.25 ± 5.90 161.94 ± 61.13 1171.22 ± 121.95
29.30 ± 4.29 140.07 ± 43.09 1170.78 ± 174.17
36.75 ± 5.75 156.21 ± 67.97 1199.10 ± 179.76
4.28 ± 1.32
5.66 ± 0.89
5.08 ± 1.51
No. of animals Creatinine (mg/dl) Urea (mg/dl) LDHplasma (U/l) LDHerythrocytes (U/l) GOT (U/l) ChE (U/l) ALP (U/l)
89.83 ± 14.67 287.37 ± 70.07 125.90 ± 41.82
90.25 ± 22.37 274.50 ± 71.71 123.00 ± 50.54
97.30 ± 14.08 289.69 ± 36.22 143.10 ± 51.53
124.72 ± 21.25a
88.83 ± 26.32b
90.47 ± 17.06b
40.32 ± 8.97a
30.63 ± 12.51ab
25.55 ± 8.11b
0.09 ± 0.02a
0.08 ± 0.03a
0.05 ± 0.02b
0.83 ± 0.44
0.85 ± 0.48
2.09 ± 1.35
47.90 ± 11.94
53.35 ± 23.67
67.53 ± 33.70
Results are expressed as mean values ± SD. A Statistics: Kruskal–Wallis and Mann–Whitney U-test. B ANOVA and Tukey test. Values in the same row not showing a common letter (a, b) are significantly different at P < 0.05.
Serum parameters are shown in Fig. 1A and B and Table 2. BDE99 caused a significant increase in GPT activity (Fig. 1A) at 1.2 mg BDE-99/kg, and a significant decrease in uric acid levels at both BDE-99 doses (Fig. 1B). Moreover, there was a tendency to enhance GOT and ALP activities at 1.2 mg BDE-99/kg. The remaining parameters in blood remained unaltered after BDE-99 administration (Table 2).
2500
A b
40
U GPT/l
The analysis of blood OS biomarkers revealed a significant increase in SOD activity (Fig. 2A), while GR activity diminished in both BDE-99 exposed groups (Fig. 2B). OS biomarkers in plasma and erythrocytes are summarized in Table 3. In erythrocytes, there was a trend to reduce CAT activity, while GPx activity was enhanced due to BDE-99 exposure. A tendency to reduce GSSG levels and GSSG/GSH ratio was also noted in rats exposed to 1.2 mg/kg of BDE-99. With respect to the remaining OS biomarkers, there were no significant differences between control and BDE-99 exposed groups. Circulating concentrations of sex steroids and thyroid hormones are shown in Fig. 3A and B and Table 4. Testosterone levels significantly decreased in both BDE-99 treated groups (Fig. 3A). Progesterone was significantly decreased in rats given 1.2 mg BDE-99/kg (Fig. 3B). A tendency to enhance TSH was also noted in the BDE-99 exposed groups, while no significant differences were observed in serum T3 and T4 (total and free) in comparison to the control group (Table 4).
30
a
20
ab
A
2000
U SOD/ g Hb
50
Results are expressed as mean values ± SD. Statistics: ANOVA.
b b
1500 1000
a 500
10
0
0 Control
BDE 0.6
Control
BDE 1.2
Groups 2
375
B a
BDE 1.2
300
B a
1.6 1.2
ab
b
0.8
U GR/g Hb
mg uric acid/dl
BDE 0.6
Groups
225 150
b
b
75
0.4 0
0
Control
Control
BDE 0.6
BDE 1.2
BDE 0.6
BDE 1.2
Groups
Groups Fig. 1. Serum activities of GPT (A) and uric acid levels (B) in adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by Kruskal–Wallis, followed by the Mann–Whitney U-test. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
Fig. 2. Antioxidant activities of SOD (A) and GR (B) in erythrocytes of adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by ANOVA, followed by the Tukey test for SOD activity, and Kruskal–Wallis, followed by the Mann–Whitney U-test, for GR activity. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211 Table 3 Blood oxidative stress biomarkers in adult rats exposed to BDE-99. Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals Erythrocytes CAT (mmol/min/g Hb)b GPx (U/g Hb)b TBARS (nmol/g Hb)b GSH (lmol/g Hb)b GSSG (lmol/g Hb)b GSSG/GSHa
10
10
10
34.52 ± 14.91
32.30 ± 4.78
29.60 ± 4.94
106.77 ± 28.12 6.11 ± 2.34 0.32 ± 0.23 2.36 ± 0.62 13.39 ± 9.62
132.87 ± 26.35 5.68 ± 1.96 0.27 ± 0.16 2.30 ± 0.66 10.88 ± 5.91
130.40 ± 29.88 4.58 ± 2.25 0.36 ± 0.24 2.00 ± 0.40 7.81 ± 3.59
26.37 ± 6.15 17.96 ± 5.60 57.09 ± 15.53 3.49 ± 1.58
24.94 ± 9.96 17.75 ± 7.86 49.55 ± 14.74 3.24 ± 0.47
19.39 ± 7.78 20.76 ± 4.41 46.07 ± 6.40 4.65 ± 2.76
Plasma TBARS (lmol/ml)b GSH (nmol/ml)b GSSG (nmol/ml)b GSSG/GSHb
Results are expressed as mean values ± SD. a Statistics: Kruskal–Wallis. b ANOVA.
Testosterone (mmol/l)
20
a
A
16 12
b
b
8 4 0 Control
BDE 0.6
BDE 1.2
Groups
Progesterone (mmol/l)
25
B a
20
ab
15
b 10 5 0 Control
BDE 0.6
BDE 1.2
Groups Fig. 3. Serum levels of testosterone (A) and progesterone (B) in adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by ANOVA, followed by the Tukey test. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
Table 4 Concentrations of serum thyroid hormones in adult rats exposed to BDE-99. Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals TSH (U/l) T3 (nmol/l) T4 (nmol/l) Free T4 (pmol/l)
10 1.44 ± 0.58 1.32 ± 0.12 51.42 ± 4.17 35.80 ± 1.67
10 1.98 ± 0.61 1.21 ± 0.08 50.02 ± 5.78 35.02 ± 2.26
10 2.07 ± 0.65 1.32 ± 0.09 50.69 ± 6.89 34.90 ± 3.48
Results are expressed as mean values ± SD. Statistics: ANOVA.
4. Discussion In the present investigation, the effects of BDE-99 exposure on biochemical parameters and hormonal function were assessed in
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adult male rats. To date, the toxic effects of BDE-99 on renal function were not well described. In the current study, urine analysis showed an increase in the protein content of adult rats orally exposed to a single dose of BDE-99. Glomerular filtrate contains a high concentration of proteins, which are mainly reabsorbed by the tubules. Most proteins are too large to pass through the filtering system of the kidney and consequently they are not present in urine under normal conditions. An increase of proteinuria is associated with renal pathology (Jarad and Miner, 2009). Data on enzyme activities in urine also facilitate the detection of renal damage (Thomsen et al., 1993; Bellés et al. 2007). The results of the present study show that BDE-99 exposure caused a decreased in LDH, GGT and NAG activities. Previous studies about hepatic toxicity were rather focused on changes in weight and size of the liver (ATSDR, 2004). Serum GOT, GPT and ALP are considered the most sensitive markers for the diagnosis of hepatotoxicity. These enzymes are predominant in cytoplasm of hepatic cells, and their excretions to blood suggest cell damage. Our current findings show that BDE-99 exposure increased those enzymatic activities (although not significantly in all cases). These results might indicate that the alteration of hepatic cell permeability was not severe, or that the period of 45 days after exposure was enough to restore initial levels (Amacher, 1998). Similar results were found in previous investigations on hepatic toxicity of PCBs in rats (Mayes et al. 1998; Kutlu et al., 2007). Nowadays, the mechanism underlying BDE-99-induced toxicity is not quite clear. We here evaluated if BDE-99 could produce OS as a possible mechanism of toxicity. This hypothesis was based on the results of previous studies demonstrating that exposure to environmental contaminants such as PCBs, caused disruptions through OS (Lehmann et al., 2007; Glauert et al., 2008). In the current study, urinary isoprostane levels were enhanced due to BDE-99 exposure, which reflects lipid peroxidation of arachidonic acid present in membranes, suggesting renal oxidative damage (Morrow et al., 1994). Furthermore, BDE-99 decreased uric acid levels in blood. Uric acid is a powerful non-enzymatic antioxidant, which acts as a scavenger of singlet oxygen and radicals (Glantzounis et al., 2005; Kuzkaya et al., 2005). The decrease in uric acid levels suggests that BDE-99 may cause OS. Blood antioxidant capacity is predominantly situated in erythrocytes. These cells neutralize ROS and protect themselves, and also the rest of tissues, from oxidative damage (Mendiratta et al., 1998; Siems et al., 2000). The antioxidant defense systems are closely interrelated. Superoxide anions are neutralized by SOD, and hydrogen peroxide is formed. In those animals with a high production of superoxide anions, a greater production of hydrogen peroxide can be expected (SOD activity can be induced) (Linares et al., 2006). We observed an increase of SOD in erythrocytes of animals exposed to BDE-99. Moreover, a decrease in GR activity was also noted. Both enzymes are considered as the first line of defense against deleterious effects of ROS. An intense production of ROS during time prolongation can exhaust enzymatic activities as we observed in GR. It has been reported that PBDEs cause developmental and reproductive effects suggesting endocrine disruptor behaviour (Darnerud, 2008). Due to the lack of information about the effects of individual PBDE congeners, we also carried out a hormonal analysis to evaluate the influence of BDE-99 on hormonal response. Gestational BDE-99 exposure produced alterations in TSH, T3 and T4 levels (Hakk et al., 2002; Kuriyama et al., 2007). In contrast, our data show that BDE-99 exposure of adult animals did not affect these hormone levels. These results suggest that the effects of BDE99 on thyroid hormones are present when exposure occurred during a critical phase of development, but not in adult animals. Moreover, the period of 45 days after exposure could have been sufficient to restore hormonal levels.
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Although BDE-99 interaction with thyroid function has been corroborated, there is a lack of information about the effects on other hormonal systems. It has been reported that exposure to BDE-99 during gestation decreased circulating sex steroids (Lilienthal et al., 2006). Notwithstanding, there is no information on the effects of these hormones in adult animals. In the present study, BDE-99 exposure to adult rats produced a decrease in testosterone and its precursor progesterone. A similar effect was also observed by other authors due to PCB exposure (Andric et al., 2000; Murugesan et al., 2008). It suggests that BDE-99 may have anti-androgenic activity. However, further investigations are needed in order to clarify if BDE-99 inhibits androgen synthesis, binds to the androgen receptor (AR) (Stoker et al., 2005), or competes for serum hormone-binding protein. Testosterone is required for the maturation of male germ cells, the production of sperm, and thus male fertility (McLachlan et al., 2002; Walker, 2009). In a recent study of our research group, we found that adult BDE-99 exposure resulted in adverse changes in the male reproductive system of rats (unpublished data). The decreased testosterone levels found in the current investigation could be responsible for the spermatid counts decrease and morphological sperm alterations observed in that recent study. It is important to relate the concentrations of BDE-99 utilized in this study with levels of PBDEs found in human populations. The doses of BDE-99 here used (0.6 and 1.2 mg/kg) corresponded to 58 and 116 times the highest level found in human adipose tissue (She et al., 2002). These are the lowest doses used in adult rats, for which toxic effects were observed. The application of a total uncertainty factor (UF) of 3000 to the oral BMDL1SD (95% lower confidence limit on the maximum likelihood estimate of the dose corresponding to a change in the mean equal to one standard deviation (SD) of the control mean) of 0.29 mg/kg in mice results in an RfD (reference dose) for BDE-99 of 0.0001 mg/kg-day, being neurobehavioral changes the critical effects (IRIS, 2008). A BMD1SD (maximum likelihood estimate of the dose corresponding to a change in the mean equal to one SD of the control mean) of 0.41 mg/kg was found in a single dose (gavage) study in mice (IRIS, 2008). On the other hand, ATSDR (2004) derived a minimal risk level (MRL) of 0.03 mg/kg/day for acute-duration oral exposure to lower PBDEs. The MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse, non-cancer health effects over a specified duration of exposure. The acute oral MRL for PBDEs is based on a NOAEL of 1 mg/kg/day for reduced serum levels of thyroid T4 hormone in fetal rats that were exposed to a commercial penta-BDE mixture on days 4–20 of gestation (Zhou et al., 2002). ATSDR also derived an MRL of 0.007 mg/kg/day for intermediate-duration oral exposure to lower brominated PBDEs. The intermediate oral MRL is based on a minimal LOAEL of 2 mg/kg/day for liver effects in rats that were exposed to a commercial penta-BDE mixture for 90 days. A chronic-duration oral MRL was not derived for lower brominated PBDEs due to insufficient data (ATSDR, 2004). Nowadays, to the best of our knowledge, no specific MRL and RfD for oral BDE-99 are available. Moreover, as above commented most experimental studies on the effects of PBDE exposure have been performed during development, with evident lack information about toxic effects in adulthood. In addition to environmental exposure, including the diet (Darnerud et al., 2001; Domingo, 2004; Domingo et al., 2006, 2008), another potentially important source of PBDE exposure is occupational. This is a clear reason to investigate the toxic effects of PBDEs in adulthood. On the other hand, the mechanism through which BDE-99 causes toxicity remains still to be elucidated. An additional gap on the knowledge of toxicity of PBDEs is that most experimental studies have focused in effects due to BDE-mixtures, but not due to individual congeners.
In summary, the results of this study demonstrates for the first time that adult exposure to low doses of BDE-99 causes biochemical alterations and also on the hormonal systems. Further investigations on the mechanism of action of BDE-99 are necessary in order to assess the human health risks of this PBDE congener. Conflict of interest The authors declare that there are no conflicts of interest. Acknowledgments The authors thank Concepción Medina, Socorro Matilla, Concepción García and Carolina Sarvisé (Hormonal Section, Laboratori Clinic ICS Camp de Tarragona, Hospital Universitari Joan XXIII) for their skilful help in the hormonal determinations. References Agarwal, A., Saleh, R.A., Bedaiwy, M.A., 2003. Role of reactive oxygen species in the pathophysiology of human reproduction. Fertil. Steril. 79, 829–843. Albina, M., Alonso, V., Linares, V., Bellés, M., Sirvent, J., Domingo, J.L., Sánchez, D.J., 2010. Effects of exposure to BDE-99 on oxidative status of liver and kidney in adult male rats. Toxicology 271, 51–56. Amacher, D.E., 1998. Serum transaminase elevations as indicators of hepatic injury following the administration of drugs. Regul. Toxicol. Pharmacol. 27, 119–130. Andric, S., Kostic, T.S., Stojilkovic, S.S., Kovacenic, R., 2000. Inhibition of rat testicular androgens by a polychlorinated biphenyl mixture Aroclor 1248. Biol. Reprod. 62, 1882–1888. ATSDR, Agency for Toxic Substances and Disease Registry, 2004. Toxicological Profile for Polybrominated Biphenyls and Polybrominated Diphenyl Ethers (PBBs and PBDEs). Department of Health and Human Services, Public Health Service, Atlanta, GA, USA. Bellés, M., Linares, V., Albina, M.L., Sirvent, J., Sánchez, D.J., Domingo, J.L., 2007. Melatonin reduces uranium-induced nephrotoxicity in rats. J. Pineal Res. 43, 87–95. Bocio, A., Llobet, J.M., Domingo, J.L., Corbella, J., Teixidó, A., Casas, C., 2003. Polybrominated diphenyl ethers (PBDEs) in foodstuffs: human exposure through the diet. J. Agric. Food Chem. 51, 3191–3195. Buege, J.A., Aust, S.D., 1978. Microsomal lipid peroxidation. Methods Enzymol. 52, 302–310. Chen, L.J., Lebetkin, E.H., Sanders, J.M., Burka, L.T., 2006. Metabolism and disposition of 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE99) following a single or repeated administration to rats or mice. Xenobiotica 36, 515–534. Cheng, J., Gu, J., Ma, J., Chen, X., Zhang, M., Wang, W., 2009. Neurobehavioral effects, redox responses and tissue distribution in rat offspring developmental exposure to BDE-99. Chemosphere 75, 963–968. Costa, L.G., Giordano, G., Tagliaferri, S., Caglieri, A., Mutti, A., 2008. Polybrominated diphenyl ether (PBDE) flame retardants: environmental contamination, human body burden and potential adverse health effects. Acta Biomed. 79, 172–183. Darnerud, P.O., 2003. Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 29, 841–853. Darnerud, P.O., 2008. Brominated flame retardants as possible endocrine disruptors. Int. J. Androl. 31, 152–160. Darnerud, P.O., Atuma, S., Aune, M., Bjerselius, R., Glynn, A., Grawé, K.P., Becker, W., 2006. Dietary intake estimations of organohalogen contaminants (dioxins, PCB, PBDE and chlorinated pesticides, e.g. DDT) based on Swedish market basket data. Food Chem. Toxicol. 44, 1597–1606. Darnerud, P.O., Eriksen, G.S., Jóhannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109, 49–68. Domingo, J.L., 2004. Human exposure to polybrominated diphenyl ethers through the diet. J. Chromatogr. A 1054, 321–326. Domingo, J.L., Bocio, A., Falcó, G., Llobet, J.M., 2006. Exposure to PBDEs and PCDEs associated with the consumption of edible marine species. Environ. Sci. Technol. 40, 4394–4399. Domingo, J.L., Cid, R.M., Castell, V., Llobet, J.M., 2008. Human exposure to PBDEs through the diet in Catalonia, Spain: temporal trend. A review of recent literature on dietary PBDE intake. Toxicology 248, 25–32. EFSA, 2006. EFSA European Food Safety Agency, Question No. EFSA-Q-2005-244. EFSA J. 328, 1–4. Eriksson, P., Viberg, H., Jakobsson, E., Orn, U., Frediksson, A., 2002. A brominated flame retardant, 2,20 ,4,40 ,5-pentabromodiphenyl ether: uptake, retention, and induction of neurobehavioural alterations in mice during a critical phase of neonatal brain development. Toxicol. Sci. 67, 98–103. Glantzounis, G.K., Tsimoyiannis, E.C., Kappas, A.M., Galaris, D.A., 2005. Uric acid and oxidative stress. Curr. Pharm. Des. 11, 4145–4151. Glauert, H.P., Tharappel, J.C., Lu, Z., Stemm, D., Banerjee, S., Chan, L.S., Lee, E.Y., Lehmler, H.J., Robertson, L.W., Spear, B.T., 2008. Role of oxidative stress in the promoting activities of PCBs. Environ. Toxicol. Pharmacol. 25, 247–250.
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211 Hakk, H., Larsen, G., Wehler, K., 2002. Tissue disposition, excretion and metabolism of 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE-99) in the male Sprague-Dawley rat. Xenobiotica 32, 369–382. Hamers, T., Kamstra, J.H., Sonneveld, E., Murk, A.J., Kester, M.H., Andersson, P.L., Legler, J., Brouwer, A., 2006. In vitro profiling of the endocrine-disrupting potency of brominated flame retardants. Toxicol. Sci. 92, 157–173. Hissin, P.J., Hilf, A., 1976. Fluorometric method for determination of oxidized and reduced glutathione in tissues. Anal. Biochem. 74, 214–226. IRIS, Integrated Risk Information System, 2008. US Environmental Protection Agency. Toxicological Review of 2,20 ,4,40 ,5-Pentabromodiphenyl Ether (BDE99) (CAS No. 60348-60-9). Available from: http://www.epa.gov/iris/. Jarad, G., Miner, J.H., 2009. Update on the glomerular filtration barrier. Curr. Opin. Nephrol. Hypertens. 18, 226–232. Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005. Polybrominated diphenyl ethers and polychlorinated biphenyls in human adipose tissue from New York. Environ. Sci. Technol. 39, 5177–5182. Kiruthiga, P.V., Shafreen, R.B., Pandian, S.K., Arun, S., Govindu, S., Devi, K.P., 2007. Protective effect of silymarin on erythrocyte haemolysate against benzo(a)pyrene and exogenous reactive oxygen species (H2O2) induced oxidative stress. Chemosphere 68, 1511–1518. Kuriyama, S.N., Talsness, C.E., Grote, K., Chahoud, I., 2005. Developmental exposure to low dose PBDE 99: effects on male fertility and neurobehavior in rat offspring. Environ. Health Perspect. 113, 149–154. Kuriyama, S.N., Wanner, A., Fidalgo-Neto, A.A., Talsness, C.E., Koerner, W., Chahoud, I., 2007. Developmental exposure to low-dose PBDE-99: Tissue distribution and thyroid hormone levels. Toxicology 242, 80–90. Kutlu, S., Colakoglu, N., Halifeoglu, I., Sandal, S., Seyran, A.D., Aydin, M., Ylmaz, B., 2007. Comparative evaluation of hepatotoxic and nephrotoxic effects of aroclors 1221 and 1254 in female rats. Cell Biochem. Funct. 25, 167–172. Kuzkaya, N., Weissmann, N., Harrison, D.G., Dikalov, S., 2005. Interactions of peroxynitrite with uric acid in the presence of ascorbate and thiols: implications for uncoupling endothelial nitric oxide synthase. Biochem. Pharmacol. 70, 343–354. Law, R.J., Allchin, C.R., de Boer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., de Wit, C.A., 2006. Levels and trends of brominated flame retardants in the European environment. Chemosphere 64, 187–208. Lehmann, D.W., Levine, J.F., McHugh Law, J., 2007. Polychlorinated biphenyl exposure causes gonadal atrophy and oxidative stress in Corbicula fluminea clams. Toxicol. Pathol. 35, 356–365. Lilienthal, H., Hack, A., Roth-Härer, A., Wichert Grande, S., Talsness, C.E., 2006. Effects of developmental exposure to 2,20 ,4,40 ,5-pentabromodiphenyl ether (PBDE-99) on sex steroids, sexual development, and sexually dimorphic behavior in rats. Environ. Health Perspect. 114, 194–201. Linares, V., Bellés, M., Albina, M.L., Sirvent, J.J., Sánchez, D.J., Domingo, J.L., 2006. Assessment of the pro-oxidant activity of uranium in kidney and testis of rats. Toxicol. Lett. 167, 152–161. Lind, Y., Darnerud, P.O., Atuma, S., Aune, M., Becker, W., Bjerselius, R., Cnattingius, S., Glynn, A., 2003. Polybrominated diphenyl ethers in breast milk from Uppsala County. Sweden Environ. Res. 93, 186–194. Martí-Cid, R., Bocio, A., Llobet, J.M., Domingo, J.L., 2007. Intake of chemical contaminants through fish and seafood consumption by children of Catalonia, Spain: health risks. Food Chem. Toxicol. 45, 1968–1974. Mayes, B.A., McConnell, E.E., Neal, B.H., Brunner, M.J., Hamilton, S.B., Sullivan, T.M., Peters, A.C., Ryan, M.J., Toft, J.D., Singer, A.W., Brown Jr., J.F., Menton, R.G., Moore, J.A., 1998. Comparative carcinogenicity in Sprague-Dawley rats of the polychlorinated biphenyl mixtures Aroclors 1016, 1242, 1254 and 1260. Toxicol. Sci. 41, 62–76. McLachlan, R.I., O’Donnell, L., Meachem, S.J., Stanton, P.G., De Kretser, D.M., Pratis, K., Robertson, D.M., 2002. Identification of specific sites of hormonal regulation in spermatogenesis in rats, monkeys, and man. Recent Prog. Horm. Res. 57, 149–179. Meironyte, D., Noren, K., 2001. Polybrominated diphenyl ethers in Swedish human milk. The follow-up study. In: Proceedings of the Second International Workshop on Brominated Flame Retardants, 14–16 May 2001, Stockholm, Sweden, pp. 303–305. Mendiratta, S., Qu, Z.C., May, J.M., 1998. Erythrocyte ascorbate recycling: antioxidant effects in blood. Free Radical Biol. Med. 24, 789–797. Morrow, J.D., Minton, T.A., Mukundan, C.R., Campbell, M.D., Zackert, W.E., Daniel, V.C., Badr, K.F., Blair, I.A., Roberts 2nd, L.J., 1994. Free radical-induced generation of isoprostanes in vivo. Evidence for the formation of D-ring and E-ring isoprostanes. J. Biol. Chem. 269, 4317–4326.
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Murugesan, P., Muthusamy, T., Balasubramanian, K., Arunakaran, J., 2008. Polychlorinated biphenyl (Aroclor 1254) inhibits testosterone biosynthesis ans antioxidant enzymes in cultures Leydig cells. Reprod. Toxicol. 25, 447–454. Perelló, G., Martí-Cid, R., Castell, V., Llobet, J.M., Domingo, J.L., 2009. Concentrations of polybrominated diphenyl ethers, hexachlorobenzene and polycyclic aromatic hydrocarbons in various foodstuffs before and after cooking. Food Chem. Toxicol. 47, 709–715. Richard, M.J., Guiraud, P., Meo, J., Favier, A., 1992. High-performance liquid chromatographic separation of malondialdehyde–thiobarbituric acid adduct in biological materials (plasma and human cells) using a commercially available reagent. J. Chromatogr. 577, 9–18. Sakai, H., Fukuda, G., Suzuki, N., Watanabe, C., Odawara, M., 2009. Falsely elevated thyroid-stimulating hormone (TSH) level due to macro-TSH. Endocr. J. 56, 435– 440. She, J., Petreas, M., Winkler, J., Visita, P., McKinney, M., Kopec, D., 2002. PBDEs in the San Francisco Bay area: measurement in harbor seal blubber and human breast adipose tissue. Chemosphere 46, 697–707. Shimizu, K., Ogawa, F., Thiele, J.J., Bae, S., Sato, S., 2007. Lipid peroxidation is enhanced in Yusho victims 35 years after accidental poisoning with polychlorinated biphenyls in Nagasaki. Jpn. J. Appl. Toxicol. 27, 195–197. Siems, W.G., Sommerburg, O., Grune, T., 2000. Erythrocyte free radical and energy metabolism. Clin. Nephrol. 53, 9–17. Sjödin, A., Jones, R.S., Focant, J.F., Lapeza, C., Wang, R.Y., McGahee 3rd, E.E., Zhang, Y., Turner, W.E., Slazyk, B., Needham, L.L., Patterson Jr., D.G., 2004. Retrospective time-trend study of polybrominated diphenyl ether and polybrominated and polychlorinated biphenyl levels in human serum from the United States. Environ. Health Perspect. 112, 654–658. Sjödin, A., Wong, L.Y., Jones, R.S., Park, A., Zhang, Y., Hodge, C., Dipietro, E., McClure, C., Turner, W., Needham, L.L., Patterson Jr., D.G., 2008. Serum concentrations of polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyl (PBB) in the United States population: 2003–2004. Environ. Sci. Technol. 42, 1377–1384. Stapleton, H.M., Kelly, S.M., Pei, R., Letcher, R.J., Gunsch, C., 2009. Metabolism of polybrominated diphenyl ethers (PBDEs) by human hepatocytes in vitro. Environ. Health Perspect. 117, 197–202. Stoker, T.E., Cooper, R.L., Lambright, C.S., Wilson, V.S., Furr, J., Gray, L.E., 2005. In vivo and in vitro anti-androgenic effects of DE-71, a commercial polybrominated diphenyl ether (PBDE) mixture. Toxicol. Appl. Pharmacol. 207, 78–88. Szabo, D.T., Richardson, V.M., Ross, D.G., Diliberto, J.J., Kodavanti, P.R., Birnbaum, L.S., 2009. Effects of perinatal PBDE exposure on hepatic phase I, phase II, phase III, and deiodinase 1 gene expression involved in thyroid hormone metabolism in male rat pups. Toxicol. Sci. 107, 27–39. Thomsen, H.S., Dorph, S., Larsen, S., Horn, T., Hemmingsen, L., Skaarup, P., Golman, K., Svendsen, O., 1993. Urine profiles and kidney histologic findings after intravenous injection of mannitol and iohexol in the degeneration phase of gentamicin nephropathy in rats. Invest. Radiol. 28, 133–141. Turyk, M.E., Persky, V.W., Imm, P., Knobeloch, L., Chatterton, R., Anderson, H.A., 2008. Hormone disruption by PBDEs in adult male sport fish consumers. Environ. Health Perspect. 116, 1635–1641. Venkataraman, P., Muthuvel, R., Krishnamoorthy, G., Arunkumar, A., Sridhar, M., Srinivasan, N., Balasubramanian, K., Aruldhas, M.M., Arunakaran, J., 2007. PCB (Aroclor 1254) enhances oxidative damage in rat brain regions: protective role of ascorbic acid. Neurotoxicology 28, 490–498. Walker, W.H., 2009. Molecular mechanisms of testosterone action in spermatogenesis. Steroids 74, 602–607. Wang, Y.F., Wang, S.L., Chen, F.A., Chao, H.A., Tsou, T.C., Shy, C.G., Päpke, O., Kuo, Y.M., Chao, H.R., 2008. Associations of polybrominated diphenyl ethers (PBDEs) in breast milk and dietary habits and demographic factors in Taiwan. Food Chem. Toxicol. 46, 1925–1932. Wheeler, C.R., Salzman, J.A., Elsayed, N.M., 1990. Automated assays for superoxide dismutase, catalase, glutathione peroxidase and glutathione reductase activity. Anal. Biochem. 184, 193–199. WHO, 2002. Global Assessment of the State-of-the-Science of Endocrine Disruptors. World Health Organization, WHO/PCS/EDC/02.2. Available from:
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Effects of BDE-99 on hormone homeostasis and biochemical parameters in adult male rats Virginia Alonso a,b, Victoria Linares a,b, Montserrat Bellés a,b, María Luisa Albina a,b, Andreu Pujol b, José L. Domingo a,*, Domènec J. Sánchez a,b a b
Laboratory of Toxicology and Environmental Health, School of Medicine, IISPV, Universitat Rovira i Virgili, Sant Llorens 21, 43201 Reus, Catalonia, Spain Physiology Unit, School of Medicine, IISPV, Universitat Rovira i Virgili, Sant Llorens 21, 43201 Reus, Catalonia, Spain
a r t i c l e
i n f o
Article history: Received 12 March 2010 Accepted 13 May 2010
Keywords: BDE-99 Adult rats Oxidative stress Endocrine disruption Hepatotoxicity Nephrotoxicity
a b s t r a c t In this study, we evaluated the effects of BDE-99 on hormone homeostasis, as well as in urinary and serum biochemical parameters of adult male rats. Animals (10 per group) received BDE-99 by gavage at single doses of 0, 0.6 and 1.2 mg/kg. Forty-five days after BDE-99 exposure, urine and serum samples were collected for hormonal and biochemical analysis. Oxidative stress (OS) markers in erythrocytes, plasma and urine were also evaluated. Urinary excretion of total protein significantly increased following BDE-99 exposure, while lactate dehydrogenase (LDH), c-glutamil transferase (GGT), and N-acetylglucosaminidase (NAG) activities significantly decreased. Liver toxicity was evidenced by elevated serum activities of the enzymes glutamic oxaloacetic transaminase (GOT), glutamic pyruvic transaminase (GPT) and alkaline phosphatase (ALP). Following BDE-99 administration, OS markers in erythrocytes showed an increase in superoxide dismutase (SOD) activity, and a reduction in glutathione reductase (GR) activity. In urine, isoprostane levels increased after BDE-99 exposure. The hormonal analysis showed a significant decrease in testosterone and progesterone levels. These results support the hypothesis that BDE-99 interacts with hormonal response. Moreover, BDE-99 administration to adult male rats showed signs of renal and hepatic toxicity. Ó 2010 Elsevier Ltd. All rights reserved.
1. Introduction Polybrominated diphenyl ethers (PBDEs), a group of brominated flame retardants formed by 209 congeners, are used as additives in plastics, textiles, foams and electronic appliances. These compounds have a high lipophilicity, are environmental persistent, and possess a notable potential of bioaccumulation. PBDEs are found in air, sediments, biota and human tissues (Darnerud et al., 2001; Law et al., 2006; Sjödin et al., 2008). For the general popula-
Abbreviations: ALP, alkaline phosphatase; ALT, alanine aminotransferase; PCBs, polychlorinated biphenyls; AST, aspartate aminotransferase; BDE-99, 2,20 ,4,40 ,5pentabromodiphenyl ether; CAT, catalase; ChE, cholinesterase; ER, estrogenic receptors; GGT, c-glutamil-transferase; GOT, glutamic oxaloacetic transaminase; GPT, glutamic pyruvic transaminase; GPx, glutathione peroxidase; GR, glutathione reductase; GSH, reduced glutathione; GSSG, oxidized glutathione; LDH, lactate dehydrogenase; NAG, N-acetylglucosaminidase; OS, oxidative stress; PBDEs, polybrominated diphenyl ethers; POCs, persistent organochlorinated compounds; ROS, reactive oxygen species; SOD, superoxide dismutase; T3, triiodothyronine hormone; T4, thyroxine hormone; TBARS, thiobarbituric acid reactive substances; TSH, thyroid-stimulating hormone. * Corresponding author. Tel.: +34 97759380; fax: +34 977 759322. E-mail address: [email protected] (J.L. Domingo). 0278-6915/$ - see front matter Ó 2010 Elsevier Ltd. All rights reserved. doi:10.1016/j.fct.2010.05.048
tion, the main routes of exposure to PBDEs, particularly the lower brominated congeners, is through the diet, as it also occurs with dioxins and PCBs. Consequently, consumption of foodstuffs such as fatty fish from contaminated sources is a major way of exposure to PBDEs (Bocio et al., 2003; Domingo, 2004; Darnerud et al., 2006; Domingo et al., 2006, 2008; Martí-Cid et al., 2007; Perelló et al., 2009). In previous studies on the levels of PBDEs in a number of foodstuffs, we analyzed individually the following six tetrathrough heptabrominated congeners: 47, 99, 100, 153, 154 and 183 (Domingo et al., 2006, 2008; Martí-Cid et al., 2007). For most food groups, congeners BDE-47 and BDE-99 showed the highest concentrations. These specific congeners were analyzed taking into account that they are the most abundant in the environment, being also the easiest of measuring (Domingo, 2004). In addition, these congeners, together with BDE-28 (triBDE) and BDE-209 (decaBDE), are those recommended by the European Food Safety Agency in monitoring programs for feed and food (EFSA, 2006). From a toxicological point of view and although limited data are currently available on these compounds, new toxicology data can be expected in the near future (EFSA, 2006). Human tissue concentrations of persistent organochlorinated compounds (POCs), such as polychlorinated biphenyls (PCBs) have
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211
decreased since their restriction in the 1970s (Meironyte and Noren 2001; Sjödin et al., 2004). In contrast, monitoring studies suggest that human tissue levels of PBDEs have increased in recent years (Lind et al., 2003; Sjödin et al., 2004; Johnson-Restrepo et al., 2005; Wang et al., 2008). Although there is strong evidence that PCBs and other POCs may cause adverse health effects in mammals (WHO, 2002), the potential toxicity of PBDEs in humans remains still to be thoroughly ascertained (Darnerud et al., 2001; Darnerud, 2003; Costa et al., 2008). Structurally, PBDEs are similar to PCBs, a different toxicity for the 209 possible PBDE congeners can be also expected. Among PBDE congeners, 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE-99) is one of the most persistent, being also detected in most environmental and food samples (Domingo, 2004; Domingo et al., 2006, 2008). To date, only data on a few experimental in vivo studies on BDE99 toxicity have been reported, while mechanistic information is also clearly lacking. BDE-99 has been found to elicit various adverse effects including endocrine disruption (Hakk et al., 2002), neurobehavioral toxicity (Eriksson et al., 2002; Cheng et al., 2009), and/or developmental and reproductive toxicity (Kuriyama et al., 2005; Lilienthal et al., 2006). However, most experimental studies have been focused on pre- and perinatal exposure, lacking information in adult animals. Because PBDEs have structural similarities to the hormones thyroxine (T4) and triiodothyronine (T3) (Hamers et al., 2006), concern has been raised regarding their effects on thyroid function. Although the mechanism of endocrine disruption of PBDEs has not been established yet, recently Turyk et al. (2008) found an association between PBDE exposure and increased T4 in adult male fish consumers. Moreover, little is known about PBDEs action on androgenic systems (Stoker et al., 2005). Studies on BDE-99 distribution in rats concluded that the major target for BDE-99 is adipose tissue followed by the liver where it is detoxified (Hakk et al., 2002; Chen et al., 2006). Some authors have reported an increase of hepatic microsomal enzymes activities in rats exposed to the penta-PBDE mixture DE-71 (Zhou et al., 2001; Szabo et al., 2009), while recent studies in human hepatocytes have suggested that the human liver will likely metabolize some BDE congeners (e.g., BDE-99) in vivo (Stapleton et al., 2009). On the other hand, toxic effects of BDE-99 on biochemical markers of renal function are not well known. Most studies in mammals are only based on observation of increased organ weight (ATSDR, 2004). Recently, we conducted in adult rats an investigation about renal effects of BDE-99 exposure. Results showed that BDE-99 exposure induced oxidative stress (OS) in kidneys with morphological alterations (Albina et al., 2010). In the present investigation, the role of OS as a mechanism for BDE-99 toxicity has been assessed. This hypothesis was based on the results of previous studies reporting that exposure to some persistent organic pollutants such as PCBs, could produce impairment through oxidative damage (Lehmann et al., 2007; Glauert et al., 2008). Although ROS serve as key signal molecules in physiological processes, they also play a role in pathological processes involving hormonal imbalances, neurodegenerative and reproductive impairments (Agarwal et al., 2003; Venkataraman et al., 2007). Erythrocytes can be regarded as circulating antioxidant carriers, reflecting exposure to ROS. They have been used as a model for the investigation of free-radical induced OS, because: (a) they are exposed to high oxygen tensions, (b) they are unable to replace damaged components, (c) the membrane lipids are composed partly of polyunsaturated fatty acid side chains, which are vulnerable to peroxidation, and (d) they have antioxidant enzyme systems (Kiruthiga et al., 2007). Taking the above into account, the purpose of the present study was to investigate the effects of BDE-99 on hormonal function of adult rats, as well as in blood and urine biomarkers.
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2. Materials and methods 2.1. Animals and chemical Adult male Sprague-Dawley rats (220–240 g) were obtained from Charles River (Barcelona, Spain). Animals were housed in plastic cages in a climate-controlled facility with a constant day–night cycle (light: 08:00–20:00 h) at a temperature of 22 ± 2 °C, and a relative humidity of 50 ± 10%. Food (Panlab rodent chow, Panlab, Barcelona) and tap water were offered ad libitum throughout the study. The use of animals and the experimental protocol were approved by the Animal Care and Use Committee of the ‘‘Rovira i Virgili” University (Tarragona, Catalonia, Spain). BDE-99 (98% pure) was purchased from LGC Standards S.L.U. (Barcelona). 2.2. Treatment After a quarantine period of 14 days, groups of 10 animals were given by gavage single doses of BDE-99 at 0.6 or 1.2 mg/kg of body weight (BDE 0.6 and BDE 1.2 groups, respectively). Rats in the control group received the vehicle only (corn oil). The choice of the BDE-99 doses was based on results of previous studies by Kuriyama et al. (2005). Forty-five days after BDE-99 exposure, rats were individually housed in plastic metabolism cages. Urines (24-h) were collected from animals of each group to determine biochemical parameters. After urine collection, rats were anesthetized by an intraperitoneal injection of ketamine–xylazine. Blood samples were collected from the vena cava to determine biochemical serum parameters and hormonal levels. Finally, rats were euthanized with an overdose of ketamine– xylazine. 2.3. Biochemical analysis Serum samples were analyzed for creatinine, urea, total protein, and uric acid concentrations, as well as for lactate dehydrogenase (LDH), cholinesterase (ChE), glutamic oxaloacetic transaminase (GOT), glutamic pyruvic transaminase (GPT) and alkaline phosphatase (ALP) activities using a Cobas Mira automatic analyzer (Roche Pharmaceuticals, Basel, Switzerland). In urine, volume, levels of creatinine, urea, total protein and uric acid, as well as LDH, c-glutamil-transferase (GGT), and N-acetylglucosaminidase (NAG) activities were also analyzed (Bellés et al., 2007). Urinary isoprostane levels (8-isoPGF2a) were determined as an indicator of lipid peroxidation (Cayman Chemical) (Shimizu et al., 2007). In erythrocytes and plasma, reduced (GSH) and oxidized (GSSG) glutathione were measured by the Hissin and Hilf method (1976), while thiobarbituric acid reactive substances (TBARS) were determined according to the Buege and Aust method (1978) modified for fluorimetric detection (Richard et al., 1992). In erythrocytes, superoxide dismutase (SOD), catalase (CAT) and glutathione reductase (GR) activities were determined according to previously described methods (Bellés et al., 2007), while glutathione peroxidase (GPx) activity was measured by the method of Wheeler et al. (1990). 2.4. Hormone analysis Serum samples were analyzed for thyroid-stimulating hormone (TSH), thyroxine (T4), triiodothyronine (T3), testosterone and progesterone concentrations using an Advia 2400 analyzer (Sakai et al., 2009). 2.5. Statistical analysis To evaluate the homogeneity of variances, the Levene test was used. When the variances of different treatment groups were homogeneous, ANOVA, and subsequently Tukey paired comparisons as a post hoc test, were applied to establish the level of significance among groups. If the variances were not homogeneous, the Kruskal–Wallis test and the Mann–Whitney U-test were used. The level of statistical significance for all tests was set at P < 0.05. All data were analyzed by means of the statistical package SPSS 15.0 (SPSS Sciences, Chicago, IL, USA).
3. Results The effects of exposure of adult male rats to BDE-99 on various urinary parameters are summarized in Table 1. Urinary excretion of total protein was significantly increased at 1.2 mg BDE-99/kg. LDH activity significantly decreased in the BDE-99 exposed groups, while GGT and NAG activities were significantly reduced at 1.2 mg BDE-99/kg. Although uric acid levels were increased by BDE-99, the differences with the control group were not statistically significant. Furthermore, there was a tendency to increase isoprostane levels in BDE-99 exposed groups. For the rest of urinary parameters no significant differences were noted.
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Table 1 Urinary parameters in adult rats treated with BDE-99.
Table 2 Serum parameters in adult rats treated with BDE-99.
Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals Urinary volume (ml/24 h)B Creatinine (mg/24 h)B Urea (mg/24 h)B Total protein (mg/24 h)A Uric acid (mg/24 h)B LDH (U/24 h/kg)B GGT (U/24 h/kg) B NAG (U/mg creatinine)B Creatinine clearance (ml/min)A Isoprostanes (ng 8isoPGF2a)B
10 11.75 ± 2.55
10 16.13 ± 7.71
10 13.86 ± 4.10
10 0.48 ± 0.07
10 0.48 ± 0.07
10 0.50 ± 0.11
16.44 ± 1.43
15.45 ± 1.81
15.22 ± 2.48
481.25 ± 143.37 29.22 ± 9.19a
398.75 ± 161.79 44.35 ± 23.58a
541.67 ± 189.41 126.30 ± 49.37b
32.25 ± 5.90 161.94 ± 61.13 1171.22 ± 121.95
29.30 ± 4.29 140.07 ± 43.09 1170.78 ± 174.17
36.75 ± 5.75 156.21 ± 67.97 1199.10 ± 179.76
4.28 ± 1.32
5.66 ± 0.89
5.08 ± 1.51
No. of animals Creatinine (mg/dl) Urea (mg/dl) LDHplasma (U/l) LDHerythrocytes (U/l) GOT (U/l) ChE (U/l) ALP (U/l)
89.83 ± 14.67 287.37 ± 70.07 125.90 ± 41.82
90.25 ± 22.37 274.50 ± 71.71 123.00 ± 50.54
97.30 ± 14.08 289.69 ± 36.22 143.10 ± 51.53
124.72 ± 21.25a
88.83 ± 26.32b
90.47 ± 17.06b
40.32 ± 8.97a
30.63 ± 12.51ab
25.55 ± 8.11b
0.09 ± 0.02a
0.08 ± 0.03a
0.05 ± 0.02b
0.83 ± 0.44
0.85 ± 0.48
2.09 ± 1.35
47.90 ± 11.94
53.35 ± 23.67
67.53 ± 33.70
Results are expressed as mean values ± SD. A Statistics: Kruskal–Wallis and Mann–Whitney U-test. B ANOVA and Tukey test. Values in the same row not showing a common letter (a, b) are significantly different at P < 0.05.
Serum parameters are shown in Fig. 1A and B and Table 2. BDE99 caused a significant increase in GPT activity (Fig. 1A) at 1.2 mg BDE-99/kg, and a significant decrease in uric acid levels at both BDE-99 doses (Fig. 1B). Moreover, there was a tendency to enhance GOT and ALP activities at 1.2 mg BDE-99/kg. The remaining parameters in blood remained unaltered after BDE-99 administration (Table 2).
2500
A b
40
U GPT/l
The analysis of blood OS biomarkers revealed a significant increase in SOD activity (Fig. 2A), while GR activity diminished in both BDE-99 exposed groups (Fig. 2B). OS biomarkers in plasma and erythrocytes are summarized in Table 3. In erythrocytes, there was a trend to reduce CAT activity, while GPx activity was enhanced due to BDE-99 exposure. A tendency to reduce GSSG levels and GSSG/GSH ratio was also noted in rats exposed to 1.2 mg/kg of BDE-99. With respect to the remaining OS biomarkers, there were no significant differences between control and BDE-99 exposed groups. Circulating concentrations of sex steroids and thyroid hormones are shown in Fig. 3A and B and Table 4. Testosterone levels significantly decreased in both BDE-99 treated groups (Fig. 3A). Progesterone was significantly decreased in rats given 1.2 mg BDE-99/kg (Fig. 3B). A tendency to enhance TSH was also noted in the BDE-99 exposed groups, while no significant differences were observed in serum T3 and T4 (total and free) in comparison to the control group (Table 4).
30
a
20
ab
A
2000
U SOD/ g Hb
50
Results are expressed as mean values ± SD. Statistics: ANOVA.
b b
1500 1000
a 500
10
0
0 Control
BDE 0.6
Control
BDE 1.2
Groups 2
375
B a
BDE 1.2
300
B a
1.6 1.2
ab
b
0.8
U GR/g Hb
mg uric acid/dl
BDE 0.6
Groups
225 150
b
b
75
0.4 0
0
Control
Control
BDE 0.6
BDE 1.2
BDE 0.6
BDE 1.2
Groups
Groups Fig. 1. Serum activities of GPT (A) and uric acid levels (B) in adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by Kruskal–Wallis, followed by the Mann–Whitney U-test. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
Fig. 2. Antioxidant activities of SOD (A) and GR (B) in erythrocytes of adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by ANOVA, followed by the Tukey test for SOD activity, and Kruskal–Wallis, followed by the Mann–Whitney U-test, for GR activity. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211 Table 3 Blood oxidative stress biomarkers in adult rats exposed to BDE-99. Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals Erythrocytes CAT (mmol/min/g Hb)b GPx (U/g Hb)b TBARS (nmol/g Hb)b GSH (lmol/g Hb)b GSSG (lmol/g Hb)b GSSG/GSHa
10
10
10
34.52 ± 14.91
32.30 ± 4.78
29.60 ± 4.94
106.77 ± 28.12 6.11 ± 2.34 0.32 ± 0.23 2.36 ± 0.62 13.39 ± 9.62
132.87 ± 26.35 5.68 ± 1.96 0.27 ± 0.16 2.30 ± 0.66 10.88 ± 5.91
130.40 ± 29.88 4.58 ± 2.25 0.36 ± 0.24 2.00 ± 0.40 7.81 ± 3.59
26.37 ± 6.15 17.96 ± 5.60 57.09 ± 15.53 3.49 ± 1.58
24.94 ± 9.96 17.75 ± 7.86 49.55 ± 14.74 3.24 ± 0.47
19.39 ± 7.78 20.76 ± 4.41 46.07 ± 6.40 4.65 ± 2.76
Plasma TBARS (lmol/ml)b GSH (nmol/ml)b GSSG (nmol/ml)b GSSG/GSHb
Results are expressed as mean values ± SD. a Statistics: Kruskal–Wallis. b ANOVA.
Testosterone (mmol/l)
20
a
A
16 12
b
b
8 4 0 Control
BDE 0.6
BDE 1.2
Groups
Progesterone (mmol/l)
25
B a
20
ab
15
b 10 5 0 Control
BDE 0.6
BDE 1.2
Groups Fig. 3. Serum levels of testosterone (A) and progesterone (B) in adult rats exposed to BDE-99 (0.6 or 1.2 mg/kg). Data are mean values ± SD (n = 10 per group). Significances were assessed by ANOVA, followed by the Tukey test. Groups not showing a common letter (a, b) are significantly different at P < 0.05.
Table 4 Concentrations of serum thyroid hormones in adult rats exposed to BDE-99. Group of exposure
Control
BDE 0.6 mg/kg bw
BDE 1.2 mg/kg bw
No. of animals TSH (U/l) T3 (nmol/l) T4 (nmol/l) Free T4 (pmol/l)
10 1.44 ± 0.58 1.32 ± 0.12 51.42 ± 4.17 35.80 ± 1.67
10 1.98 ± 0.61 1.21 ± 0.08 50.02 ± 5.78 35.02 ± 2.26
10 2.07 ± 0.65 1.32 ± 0.09 50.69 ± 6.89 34.90 ± 3.48
Results are expressed as mean values ± SD. Statistics: ANOVA.
4. Discussion In the present investigation, the effects of BDE-99 exposure on biochemical parameters and hormonal function were assessed in
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adult male rats. To date, the toxic effects of BDE-99 on renal function were not well described. In the current study, urine analysis showed an increase in the protein content of adult rats orally exposed to a single dose of BDE-99. Glomerular filtrate contains a high concentration of proteins, which are mainly reabsorbed by the tubules. Most proteins are too large to pass through the filtering system of the kidney and consequently they are not present in urine under normal conditions. An increase of proteinuria is associated with renal pathology (Jarad and Miner, 2009). Data on enzyme activities in urine also facilitate the detection of renal damage (Thomsen et al., 1993; Bellés et al. 2007). The results of the present study show that BDE-99 exposure caused a decreased in LDH, GGT and NAG activities. Previous studies about hepatic toxicity were rather focused on changes in weight and size of the liver (ATSDR, 2004). Serum GOT, GPT and ALP are considered the most sensitive markers for the diagnosis of hepatotoxicity. These enzymes are predominant in cytoplasm of hepatic cells, and their excretions to blood suggest cell damage. Our current findings show that BDE-99 exposure increased those enzymatic activities (although not significantly in all cases). These results might indicate that the alteration of hepatic cell permeability was not severe, or that the period of 45 days after exposure was enough to restore initial levels (Amacher, 1998). Similar results were found in previous investigations on hepatic toxicity of PCBs in rats (Mayes et al. 1998; Kutlu et al., 2007). Nowadays, the mechanism underlying BDE-99-induced toxicity is not quite clear. We here evaluated if BDE-99 could produce OS as a possible mechanism of toxicity. This hypothesis was based on the results of previous studies demonstrating that exposure to environmental contaminants such as PCBs, caused disruptions through OS (Lehmann et al., 2007; Glauert et al., 2008). In the current study, urinary isoprostane levels were enhanced due to BDE-99 exposure, which reflects lipid peroxidation of arachidonic acid present in membranes, suggesting renal oxidative damage (Morrow et al., 1994). Furthermore, BDE-99 decreased uric acid levels in blood. Uric acid is a powerful non-enzymatic antioxidant, which acts as a scavenger of singlet oxygen and radicals (Glantzounis et al., 2005; Kuzkaya et al., 2005). The decrease in uric acid levels suggests that BDE-99 may cause OS. Blood antioxidant capacity is predominantly situated in erythrocytes. These cells neutralize ROS and protect themselves, and also the rest of tissues, from oxidative damage (Mendiratta et al., 1998; Siems et al., 2000). The antioxidant defense systems are closely interrelated. Superoxide anions are neutralized by SOD, and hydrogen peroxide is formed. In those animals with a high production of superoxide anions, a greater production of hydrogen peroxide can be expected (SOD activity can be induced) (Linares et al., 2006). We observed an increase of SOD in erythrocytes of animals exposed to BDE-99. Moreover, a decrease in GR activity was also noted. Both enzymes are considered as the first line of defense against deleterious effects of ROS. An intense production of ROS during time prolongation can exhaust enzymatic activities as we observed in GR. It has been reported that PBDEs cause developmental and reproductive effects suggesting endocrine disruptor behaviour (Darnerud, 2008). Due to the lack of information about the effects of individual PBDE congeners, we also carried out a hormonal analysis to evaluate the influence of BDE-99 on hormonal response. Gestational BDE-99 exposure produced alterations in TSH, T3 and T4 levels (Hakk et al., 2002; Kuriyama et al., 2007). In contrast, our data show that BDE-99 exposure of adult animals did not affect these hormone levels. These results suggest that the effects of BDE99 on thyroid hormones are present when exposure occurred during a critical phase of development, but not in adult animals. Moreover, the period of 45 days after exposure could have been sufficient to restore hormonal levels.
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Although BDE-99 interaction with thyroid function has been corroborated, there is a lack of information about the effects on other hormonal systems. It has been reported that exposure to BDE-99 during gestation decreased circulating sex steroids (Lilienthal et al., 2006). Notwithstanding, there is no information on the effects of these hormones in adult animals. In the present study, BDE-99 exposure to adult rats produced a decrease in testosterone and its precursor progesterone. A similar effect was also observed by other authors due to PCB exposure (Andric et al., 2000; Murugesan et al., 2008). It suggests that BDE-99 may have anti-androgenic activity. However, further investigations are needed in order to clarify if BDE-99 inhibits androgen synthesis, binds to the androgen receptor (AR) (Stoker et al., 2005), or competes for serum hormone-binding protein. Testosterone is required for the maturation of male germ cells, the production of sperm, and thus male fertility (McLachlan et al., 2002; Walker, 2009). In a recent study of our research group, we found that adult BDE-99 exposure resulted in adverse changes in the male reproductive system of rats (unpublished data). The decreased testosterone levels found in the current investigation could be responsible for the spermatid counts decrease and morphological sperm alterations observed in that recent study. It is important to relate the concentrations of BDE-99 utilized in this study with levels of PBDEs found in human populations. The doses of BDE-99 here used (0.6 and 1.2 mg/kg) corresponded to 58 and 116 times the highest level found in human adipose tissue (She et al., 2002). These are the lowest doses used in adult rats, for which toxic effects were observed. The application of a total uncertainty factor (UF) of 3000 to the oral BMDL1SD (95% lower confidence limit on the maximum likelihood estimate of the dose corresponding to a change in the mean equal to one standard deviation (SD) of the control mean) of 0.29 mg/kg in mice results in an RfD (reference dose) for BDE-99 of 0.0001 mg/kg-day, being neurobehavioral changes the critical effects (IRIS, 2008). A BMD1SD (maximum likelihood estimate of the dose corresponding to a change in the mean equal to one SD of the control mean) of 0.41 mg/kg was found in a single dose (gavage) study in mice (IRIS, 2008). On the other hand, ATSDR (2004) derived a minimal risk level (MRL) of 0.03 mg/kg/day for acute-duration oral exposure to lower PBDEs. The MRL is an estimate of the daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse, non-cancer health effects over a specified duration of exposure. The acute oral MRL for PBDEs is based on a NOAEL of 1 mg/kg/day for reduced serum levels of thyroid T4 hormone in fetal rats that were exposed to a commercial penta-BDE mixture on days 4–20 of gestation (Zhou et al., 2002). ATSDR also derived an MRL of 0.007 mg/kg/day for intermediate-duration oral exposure to lower brominated PBDEs. The intermediate oral MRL is based on a minimal LOAEL of 2 mg/kg/day for liver effects in rats that were exposed to a commercial penta-BDE mixture for 90 days. A chronic-duration oral MRL was not derived for lower brominated PBDEs due to insufficient data (ATSDR, 2004). Nowadays, to the best of our knowledge, no specific MRL and RfD for oral BDE-99 are available. Moreover, as above commented most experimental studies on the effects of PBDE exposure have been performed during development, with evident lack information about toxic effects in adulthood. In addition to environmental exposure, including the diet (Darnerud et al., 2001; Domingo, 2004; Domingo et al., 2006, 2008), another potentially important source of PBDE exposure is occupational. This is a clear reason to investigate the toxic effects of PBDEs in adulthood. On the other hand, the mechanism through which BDE-99 causes toxicity remains still to be elucidated. An additional gap on the knowledge of toxicity of PBDEs is that most experimental studies have focused in effects due to BDE-mixtures, but not due to individual congeners.
In summary, the results of this study demonstrates for the first time that adult exposure to low doses of BDE-99 causes biochemical alterations and also on the hormonal systems. Further investigations on the mechanism of action of BDE-99 are necessary in order to assess the human health risks of this PBDE congener. Conflict of interest The authors declare that there are no conflicts of interest. Acknowledgments The authors thank Concepción Medina, Socorro Matilla, Concepción García and Carolina Sarvisé (Hormonal Section, Laboratori Clinic ICS Camp de Tarragona, Hospital Universitari Joan XXIII) for their skilful help in the hormonal determinations. References Agarwal, A., Saleh, R.A., Bedaiwy, M.A., 2003. Role of reactive oxygen species in the pathophysiology of human reproduction. Fertil. Steril. 79, 829–843. Albina, M., Alonso, V., Linares, V., Bellés, M., Sirvent, J., Domingo, J.L., Sánchez, D.J., 2010. Effects of exposure to BDE-99 on oxidative status of liver and kidney in adult male rats. Toxicology 271, 51–56. Amacher, D.E., 1998. Serum transaminase elevations as indicators of hepatic injury following the administration of drugs. Regul. Toxicol. Pharmacol. 27, 119–130. Andric, S., Kostic, T.S., Stojilkovic, S.S., Kovacenic, R., 2000. Inhibition of rat testicular androgens by a polychlorinated biphenyl mixture Aroclor 1248. Biol. Reprod. 62, 1882–1888. ATSDR, Agency for Toxic Substances and Disease Registry, 2004. Toxicological Profile for Polybrominated Biphenyls and Polybrominated Diphenyl Ethers (PBBs and PBDEs). Department of Health and Human Services, Public Health Service, Atlanta, GA, USA. Bellés, M., Linares, V., Albina, M.L., Sirvent, J., Sánchez, D.J., Domingo, J.L., 2007. Melatonin reduces uranium-induced nephrotoxicity in rats. J. Pineal Res. 43, 87–95. Bocio, A., Llobet, J.M., Domingo, J.L., Corbella, J., Teixidó, A., Casas, C., 2003. Polybrominated diphenyl ethers (PBDEs) in foodstuffs: human exposure through the diet. J. Agric. Food Chem. 51, 3191–3195. Buege, J.A., Aust, S.D., 1978. Microsomal lipid peroxidation. Methods Enzymol. 52, 302–310. Chen, L.J., Lebetkin, E.H., Sanders, J.M., Burka, L.T., 2006. Metabolism and disposition of 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE99) following a single or repeated administration to rats or mice. Xenobiotica 36, 515–534. Cheng, J., Gu, J., Ma, J., Chen, X., Zhang, M., Wang, W., 2009. Neurobehavioral effects, redox responses and tissue distribution in rat offspring developmental exposure to BDE-99. Chemosphere 75, 963–968. Costa, L.G., Giordano, G., Tagliaferri, S., Caglieri, A., Mutti, A., 2008. Polybrominated diphenyl ether (PBDE) flame retardants: environmental contamination, human body burden and potential adverse health effects. Acta Biomed. 79, 172–183. Darnerud, P.O., 2003. Toxic effects of brominated flame retardants in man and in wildlife. Environ. Int. 29, 841–853. Darnerud, P.O., 2008. Brominated flame retardants as possible endocrine disruptors. Int. J. Androl. 31, 152–160. Darnerud, P.O., Atuma, S., Aune, M., Bjerselius, R., Glynn, A., Grawé, K.P., Becker, W., 2006. Dietary intake estimations of organohalogen contaminants (dioxins, PCB, PBDE and chlorinated pesticides, e.g. DDT) based on Swedish market basket data. Food Chem. Toxicol. 44, 1597–1606. Darnerud, P.O., Eriksen, G.S., Jóhannesson, T., Larsen, P.B., Viluksela, M., 2001. Polybrominated diphenyl ethers: occurrence, dietary exposure, and toxicology. Environ. Health Perspect. 109, 49–68. Domingo, J.L., 2004. Human exposure to polybrominated diphenyl ethers through the diet. J. Chromatogr. A 1054, 321–326. Domingo, J.L., Bocio, A., Falcó, G., Llobet, J.M., 2006. Exposure to PBDEs and PCDEs associated with the consumption of edible marine species. Environ. Sci. Technol. 40, 4394–4399. Domingo, J.L., Cid, R.M., Castell, V., Llobet, J.M., 2008. Human exposure to PBDEs through the diet in Catalonia, Spain: temporal trend. A review of recent literature on dietary PBDE intake. Toxicology 248, 25–32. EFSA, 2006. EFSA European Food Safety Agency, Question No. EFSA-Q-2005-244. EFSA J. 328, 1–4. Eriksson, P., Viberg, H., Jakobsson, E., Orn, U., Frediksson, A., 2002. A brominated flame retardant, 2,20 ,4,40 ,5-pentabromodiphenyl ether: uptake, retention, and induction of neurobehavioural alterations in mice during a critical phase of neonatal brain development. Toxicol. Sci. 67, 98–103. Glantzounis, G.K., Tsimoyiannis, E.C., Kappas, A.M., Galaris, D.A., 2005. Uric acid and oxidative stress. Curr. Pharm. Des. 11, 4145–4151. Glauert, H.P., Tharappel, J.C., Lu, Z., Stemm, D., Banerjee, S., Chan, L.S., Lee, E.Y., Lehmler, H.J., Robertson, L.W., Spear, B.T., 2008. Role of oxidative stress in the promoting activities of PCBs. Environ. Toxicol. Pharmacol. 25, 247–250.
V. Alonso et al. / Food and Chemical Toxicology 48 (2010) 2206–2211 Hakk, H., Larsen, G., Wehler, K., 2002. Tissue disposition, excretion and metabolism of 2,20 ,4,40 ,5-pentabromodiphenyl ether (BDE-99) in the male Sprague-Dawley rat. Xenobiotica 32, 369–382. Hamers, T., Kamstra, J.H., Sonneveld, E., Murk, A.J., Kester, M.H., Andersson, P.L., Legler, J., Brouwer, A., 2006. In vitro profiling of the endocrine-disrupting potency of brominated flame retardants. Toxicol. Sci. 92, 157–173. Hissin, P.J., Hilf, A., 1976. Fluorometric method for determination of oxidized and reduced glutathione in tissues. Anal. Biochem. 74, 214–226. IRIS, Integrated Risk Information System, 2008. US Environmental Protection Agency. Toxicological Review of 2,20 ,4,40 ,5-Pentabromodiphenyl Ether (BDE99) (CAS No. 60348-60-9). Available from: http://www.epa.gov/iris/. Jarad, G., Miner, J.H., 2009. Update on the glomerular filtration barrier. Curr. Opin. Nephrol. Hypertens. 18, 226–232. Johnson-Restrepo, B., Kannan, K., Rapaport, D.P., Rodan, B.D., 2005. Polybrominated diphenyl ethers and polychlorinated biphenyls in human adipose tissue from New York. Environ. Sci. Technol. 39, 5177–5182. Kiruthiga, P.V., Shafreen, R.B., Pandian, S.K., Arun, S., Govindu, S., Devi, K.P., 2007. Protective effect of silymarin on erythrocyte haemolysate against benzo(a)pyrene and exogenous reactive oxygen species (H2O2) induced oxidative stress. Chemosphere 68, 1511–1518. Kuriyama, S.N., Talsness, C.E., Grote, K., Chahoud, I., 2005. Developmental exposure to low dose PBDE 99: effects on male fertility and neurobehavior in rat offspring. Environ. Health Perspect. 113, 149–154. Kuriyama, S.N., Wanner, A., Fidalgo-Neto, A.A., Talsness, C.E., Koerner, W., Chahoud, I., 2007. Developmental exposure to low-dose PBDE-99: Tissue distribution and thyroid hormone levels. Toxicology 242, 80–90. Kutlu, S., Colakoglu, N., Halifeoglu, I., Sandal, S., Seyran, A.D., Aydin, M., Ylmaz, B., 2007. Comparative evaluation of hepatotoxic and nephrotoxic effects of aroclors 1221 and 1254 in female rats. Cell Biochem. Funct. 25, 167–172. Kuzkaya, N., Weissmann, N., Harrison, D.G., Dikalov, S., 2005. Interactions of peroxynitrite with uric acid in the presence of ascorbate and thiols: implications for uncoupling endothelial nitric oxide synthase. Biochem. Pharmacol. 70, 343–354. Law, R.J., Allchin, C.R., de Boer, J., Covaci, A., Herzke, D., Lepom, P., Morris, S., Tronczynski, J., de Wit, C.A., 2006. Levels and trends of brominated flame retardants in the European environment. Chemosphere 64, 187–208. Lehmann, D.W., Levine, J.F., McHugh Law, J., 2007. Polychlorinated biphenyl exposure causes gonadal atrophy and oxidative stress in Corbicula fluminea clams. Toxicol. Pathol. 35, 356–365. Lilienthal, H., Hack, A., Roth-Härer, A., Wichert Grande, S., Talsness, C.E., 2006. Effects of developmental exposure to 2,20 ,4,40 ,5-pentabromodiphenyl ether (PBDE-99) on sex steroids, sexual development, and sexually dimorphic behavior in rats. Environ. Health Perspect. 114, 194–201. Linares, V., Bellés, M., Albina, M.L., Sirvent, J.J., Sánchez, D.J., Domingo, J.L., 2006. Assessment of the pro-oxidant activity of uranium in kidney and testis of rats. Toxicol. Lett. 167, 152–161. Lind, Y., Darnerud, P.O., Atuma, S., Aune, M., Becker, W., Bjerselius, R., Cnattingius, S., Glynn, A., 2003. Polybrominated diphenyl ethers in breast milk from Uppsala County. Sweden Environ. Res. 93, 186–194. Martí-Cid, R., Bocio, A., Llobet, J.M., Domingo, J.L., 2007. Intake of chemical contaminants through fish and seafood consumption by children of Catalonia, Spain: health risks. Food Chem. Toxicol. 45, 1968–1974. Mayes, B.A., McConnell, E.E., Neal, B.H., Brunner, M.J., Hamilton, S.B., Sullivan, T.M., Peters, A.C., Ryan, M.J., Toft, J.D., Singer, A.W., Brown Jr., J.F., Menton, R.G., Moore, J.A., 1998. Comparative carcinogenicity in Sprague-Dawley rats of the polychlorinated biphenyl mixtures Aroclors 1016, 1242, 1254 and 1260. Toxicol. Sci. 41, 62–76. McLachlan, R.I., O’Donnell, L., Meachem, S.J., Stanton, P.G., De Kretser, D.M., Pratis, K., Robertson, D.M., 2002. Identification of specific sites of hormonal regulation in spermatogenesis in rats, monkeys, and man. Recent Prog. Horm. Res. 57, 149–179. Meironyte, D., Noren, K., 2001. Polybrominated diphenyl ethers in Swedish human milk. The follow-up study. In: Proceedings of the Second International Workshop on Brominated Flame Retardants, 14–16 May 2001, Stockholm, Sweden, pp. 303–305. Mendiratta, S., Qu, Z.C., May, J.M., 1998. Erythrocyte ascorbate recycling: antioxidant effects in blood. Free Radical Biol. Med. 24, 789–797. Morrow, J.D., Minton, T.A., Mukundan, C.R., Campbell, M.D., Zackert, W.E., Daniel, V.C., Badr, K.F., Blair, I.A., Roberts 2nd, L.J., 1994. Free radical-induced generation of isoprostanes in vivo. Evidence for the formation of D-ring and E-ring isoprostanes. J. Biol. Chem. 269, 4317–4326.
2211
Murugesan, P., Muthusamy, T., Balasubramanian, K., Arunakaran, J., 2008. Polychlorinated biphenyl (Aroclor 1254) inhibits testosterone biosynthesis ans antioxidant enzymes in cultures Leydig cells. Reprod. Toxicol. 25, 447–454. Perelló, G., Martí-Cid, R., Castell, V., Llobet, J.M., Domingo, J.L., 2009. Concentrations of polybrominated diphenyl ethers, hexachlorobenzene and polycyclic aromatic hydrocarbons in various foodstuffs before and after cooking. Food Chem. Toxicol. 47, 709–715. Richard, M.J., Guiraud, P., Meo, J., Favier, A., 1992. High-performance liquid chromatographic separation of malondialdehyde–thiobarbituric acid adduct in biological materials (plasma and human cells) using a commercially available reagent. J. Chromatogr. 577, 9–18. Sakai, H., Fukuda, G., Suzuki, N., Watanabe, C., Odawara, M., 2009. Falsely elevated thyroid-stimulating hormone (TSH) level due to macro-TSH. Endocr. J. 56, 435– 440. She, J., Petreas, M., Winkler, J., Visita, P., McKinney, M., Kopec, D., 2002. PBDEs in the San Francisco Bay area: measurement in harbor seal blubber and human breast adipose tissue. Chemosphere 46, 697–707. Shimizu, K., Ogawa, F., Thiele, J.J., Bae, S., Sato, S., 2007. Lipid peroxidation is enhanced in Yusho victims 35 years after accidental poisoning with polychlorinated biphenyls in Nagasaki. Jpn. J. Appl. Toxicol. 27, 195–197. Siems, W.G., Sommerburg, O., Grune, T., 2000. Erythrocyte free radical and energy metabolism. Clin. Nephrol. 53, 9–17. Sjödin, A., Jones, R.S., Focant, J.F., Lapeza, C., Wang, R.Y., McGahee 3rd, E.E., Zhang, Y., Turner, W.E., Slazyk, B., Needham, L.L., Patterson Jr., D.G., 2004. Retrospective time-trend study of polybrominated diphenyl ether and polybrominated and polychlorinated biphenyl levels in human serum from the United States. Environ. Health Perspect. 112, 654–658. Sjödin, A., Wong, L.Y., Jones, R.S., Park, A., Zhang, Y., Hodge, C., Dipietro, E., McClure, C., Turner, W., Needham, L.L., Patterson Jr., D.G., 2008. Serum concentrations of polybrominated diphenyl ethers (PBDEs) and polybrominated biphenyl (PBB) in the United States population: 2003–2004. Environ. Sci. Technol. 42, 1377–1384. Stapleton, H.M., Kelly, S.M., Pei, R., Letcher, R.J., Gunsch, C., 2009. Metabolism of polybrominated diphenyl ethers (PBDEs) by human hepatocytes in vitro. Environ. Health Perspect. 117, 197–202. Stoker, T.E., Cooper, R.L., Lambright, C.S., Wilson, V.S., Furr, J., Gray, L.E., 2005. In vivo and in vitro anti-androgenic effects of DE-71, a commercial polybrominated diphenyl ether (PBDE) mixture. Toxicol. Appl. Pharmacol. 207, 78–88. Szabo, D.T., Richardson, V.M., Ross, D.G., Diliberto, J.J., Kodavanti, P.R., Birnbaum, L.S., 2009. Effects of perinatal PBDE exposure on hepatic phase I, phase II, phase III, and deiodinase 1 gene expression involved in thyroid hormone metabolism in male rat pups. Toxicol. Sci. 107, 27–39. Thomsen, H.S., Dorph, S., Larsen, S., Horn, T., Hemmingsen, L., Skaarup, P., Golman, K., Svendsen, O., 1993. Urine profiles and kidney histologic findings after intravenous injection of mannitol and iohexol in the degeneration phase of gentamicin nephropathy in rats. Invest. Radiol. 28, 133–141. Turyk, M.E., Persky, V.W., Imm, P., Knobeloch, L., Chatterton, R., Anderson, H.A., 2008. Hormone disruption by PBDEs in adult male sport fish consumers. Environ. Health Perspect. 116, 1635–1641. Venkataraman, P., Muthuvel, R., Krishnamoorthy, G., Arunkumar, A., Sridhar, M., Srinivasan, N., Balasubramanian, K., Aruldhas, M.M., Arunakaran, J., 2007. PCB (Aroclor 1254) enhances oxidative damage in rat brain regions: protective role of ascorbic acid. Neurotoxicology 28, 490–498. Walker, W.H., 2009. Molecular mechanisms of testosterone action in spermatogenesis. Steroids 74, 602–607. Wang, Y.F., Wang, S.L., Chen, F.A., Chao, H.A., Tsou, T.C., Shy, C.G., Päpke, O., Kuo, Y.M., Chao, H.R., 2008. Associations of polybrominated diphenyl ethers (PBDEs) in breast milk and dietary habits and demographic factors in Taiwan. Food Chem. Toxicol. 46, 1925–1932. Wheeler, C.R., Salzman, J.A., Elsayed, N.M., 1990. Automated assays for superoxide dismutase, catalase, glutathione peroxidase and glutathione reductase activity. Anal. Biochem. 184, 193–199. WHO, 2002. Global Assessment of the State-of-the-Science of Endocrine Disruptors. World Health Organization, WHO/PCS/EDC/02.2. Available from: