Environmental Pollution 208 (2016) 395e403
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Reliability of stable Pb isotopes to identify Pb sources and verifying biological fractionation of Pb isotopes in goats and chickens Hokuto Nakata a, 1, Shouta M.M. Nakayama a, 1, John Yabe b, Allan Liazambi c, Hazuki Mizukawa d, Wageh Sobhy Darwish a, e, Yoshinori Ikenaka a, f, Mayumi Ishizuka a, * a
Laboratory of Toxicology, Department of Environmental Veterinary Sciences, Graduate School of Veterinary Medicine, Hokkaido University, Kita 18 Nishi 9, Kita-ku, Sapporo 060-0818, Japan The University of Zambia, School of Veterinary Medicine, P.O. Box 32379, Lusaka, Zambia c Central Province Veterinary Ofﬁce, Kabwe, Zambia d Department of Environmental Veterinary Sciences, Graduate School of Veterinary Medicine, Hokkaido University, Kita 18 Nishi 9, Kita-ku, Sapporo 060-0818, Japan e Food Control Department, Faculty of Veterinary Medicine, Zagazig University, Zagazig, Egypt f Water Research Group, School of Environmental Sciences and Development, North-West University, South Africa b
a r t i c l e i n f o
a b s t r a c t
Article history: Received 30 July 2015 Received in revised form 4 October 2015 Accepted 5 October 2015 Available online 6 November 2015
Stable Pb isotope ratios (Pb-IRs) have been recognized as an efﬁcient tool for identifying sources. This study carried out at Kabwe mining area, Zambia, to elucidate the presence or absence of Pb isotope fractionation in goat and chicken, to evaluate the reliability of identifying Pb pollution sources via analysis of Pb-IRs, and to assess whether a threshold for blood Pb levels (Pb-B) for biological fractionation was present. The variation of Pb-IRs in goat decreased with an increase in Pb-B and were ﬁxed at certain values close to those of the dominant source of Pb exposure at Pb-B > 5 mg/dL. However, chickens did not show a clear relationship for Pb-IRs against Pb-B, or a fractionation threshold. Given these, the biological fractionation of Pb isotopes should not occur in chickens but in goats, and the threshold for triggering biological fractionation is at around 5 mg/dL of Pb-B in goats. © 2015 Elsevier Ltd. All rights reserved.
Keywords: Biological fractionation Chicken Goat Pb pollution source Stable Pb isotopes
1. Introduction Among metals, lead (Pb) possesses a particularly elevated anthropogenic enrichment factor (Lantzy and Mackenzie, 1979). Widespread pollution has frequently been recorded in regions with long histories of mining and smelting, where high levels of metals contaminate water, soil, sediment, and vegetation (Razo et al., 2004; Ettler et al., 2005; Hudson-Edwards et al., 2008). Nowadays, Pb is not only a local pollutant, but is also a pollutant on a global scale due to its volatile character of Pb (Charalampides and Manoliadis, 2002). The primary sources of Pb exposure, in addition to smelters, are battery recycling, electronics, paint, traditional remedies, and leaded gasoline (Meyer et al., 2008). For animals, Pb is a non-essential and toxic metal. Even at low doses, Pb leads to
* Corresponding author. E-mail address: [email protected]
(M. Ishizuka). 1 Both authors equally contributed to this study. http://dx.doi.org/10.1016/j.envpol.2015.10.006 0269-7491/© 2015 Elsevier Ltd. All rights reserved.
neurotoxicity in humans, especially in children, due to its ability to compete with calcium (Ca2þ) in nerve function (Crosby, 1998). A recent review by Yabe et al. (2010) remarked that toxic metal contamination of the environment and livestock has reached unprecedented levels over the past decade, and that human exposure to toxic metals has become a critical component of the health risk on the African continent. This grave warning was tragically borne out by the Pb poisoning disaster in Zamfara, Nigeria, in which more than 160 people died, mostly children under the age of ﬁve (Blacksmith Institute, 2010). Given these factors, it is necessary not only to consider total concentrations and the chemical/mineralogical position of Pb, but also to precisely determine how multiple sources contribute Pb to the environment. Pb is present in the environment as four main stable isotopes: 208Pb, 207Pb, 206Pb, and 204Pb. Among these isotopes, only 204Pb has no known radiogenic parent, and its abundance in the Earth's crust does not change with time. In contrast, 206 Pb, 207Pb, and 208Pb are radiogenic isotopes, and are the products of the radioactive decay of 238U, 235U, and 232Th, respectively.
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Hence, unlike many other elements whose isotopic abundances have been ﬁxed during cosmological time, the abundance of Pb isotopes in a sample depends strictly on the concentrations of primordial Pb, uranium (U), and thorium (Th) isotopes, and the length of their decay processes (Keinonen, 1992; Sangster et al., 2000). The isotopic composition of Pb can be expressed in several ways. For instance, in the environmental sciences, the Pb isotopic composition is broadly expressed as ratios, with 208Pb/206Pb and 207 Pb/206Pb being the most preferred because these ratios can be measured precisely and analytically, and the abundances of these isotopes are relatively signiﬁcant. Additionally, it should be emphasized that the isotopic composition of Pb is not affected to a measurable extent by physico-chemical fractionation processes. Stable Pb isotope ratios (Pb-IRs), therefore, can be used as natural tracers, and serve as an efﬁcient tool for identifying sources and € fer and Rosman, 2001; Veysseyre pathways of Pb pollution (Bollho and Bollho, 2001; Charalampides and Manoliadis, 2002). Recently Pb-IRs have been generally used to identify the primary sources of Pb in air and aerosols (Chow and Johnstone, 1965; Chow and Earl, €ppel, 2000), 1972), soil (Gulson et al., 1981; Hansmann and Ko sediments (Shirahata et al., 1980; Petit et al., 1984). Moreover, sources of Pb exposure for wild birds (Scheuhammer and Templeton, 1998) and sea otters (Enhydra lutris) (Smith et al., 1990) have been identiﬁed using Pb-IRs analysis. Pb-IRs have been widely assumed to not be as fractionized in biological systems as in the environment (Rabinowitz and Wetherill, 1972; Carlson, 1996). Nevertheless, some earlier studies revealed large differences in the Pb isotopic composition among biological samples within humans (Smith et al., 1996) and within rats (Rattus norvegicus) (Wu et al., 2012; Liu et al., 2014). Studies on the distribution of each of the stable Pb isotopes in the living body are still limited. It is thus necessary to clarify whether biological fractionation of Pb isotopes occurs in the body or not in order to verify the accuracy of identifying Pb sources by Pb-IRs analysis. In fact, isotopes of light elements, such as Li (Stokes et al., 1982), C (Tieszen et al., 1983; Gannes et al., 1998), and N (Gannes et al., 1998), have been recognized to be fractionized in biological systems. Isotopic variations of heavier elements, such as Zn (Stenberg et al., 2004), Fe (Walczyk and von Blanckenburg, 2002), and Hg (Epov et al., 2008) in biological systems have also been found in recent years. If isotope fractionation occurs in the body, it should then be determined which tissues most accurately reﬂect the Pb pollution source and should thus be used to identify the source. As such, the present study was designed to elucidate the presence or absence of Pb isotope fractionation in biological samples, and to evaluate the reliability of identifying Pb pollution sources via analysis of Pb-IRs. Furthermore, Wu et al. (2012) and Liu et al. (2014) have suggested the existence of a threshold for blood Pb levels (Pb-B) for biological fractionation in rats. The current study intended to assess whether such threshold for biological fractionation was present for goats (Capra hircus) and chickens (Gallus gallus) because these animals are reared and consumed worldwide. For these purposes, Kabwe, the provincial capital of Zambia's central province, was selected as a study site because it has been well studied, with ﬁndings of high levels of Pb accumulation in the environment such as soil (Ikenaka et al., 2010; Nakayama et al., 2011), wild rats (Rattus rattus) (Nakayama et al., 2011, 2013), livestock such as cattle and chickens (Yabe et al., 2011, 2013; Ikenaka et al., 2012), and children (Yabe et al., 2015). These studies considered Kabwe to have only one origin for Pb, and so the current study can be considered as a semi-ﬁeld study on constant Pb exposure. In addition, goats and chickens were chosen as the exposed animals because they are widely bred and consumed in developing countries, including those in Africa. As the Pb exposure route in Kabwe is still unclear, in contrast to the many earlier
studies focused merely on the extent of contamination, the current study also aims to identify the Pb source in Kabwe. 2. Materials and methods 2.1. Sampling of animals and environmental samples This study was conducted in Kabwe, which has no Pb batteryrelated facilities and is situated 130 km north of Lusaka, the capital and largest city of Zambia (supporting information Fig. S1), and in Chongwe, which is next to Lusaka. The ore in Kabwe contained sphalerite (ZnS) with Cd, galena (PbS), briarite [Cu2(Fe,Zn)GeS4], and mimetite [Pb5(AsO4)3Cl] (Kamona and Friedrich, 2007). The sampling sites were accurately located using a global positioning system (GPS), and are shown in supporting information Fig. S1 and Table S1. Details on samples are also shown in supporting information of Materials and Methods section. 2.2. Sample preparation and analysis of element concentrations All laboratory materials and instruments used in the heavy metal analysis were washed with 2% nitric acid (HNO3) and rinsed at least twice with distilled water. We conﬁrmed that there was no metal contamination through the analytical procedures using the regent (digestion) blank measurement. Samples of approximately 0.5 g of liver, kidney, lung, spleen, brain, muscle, heart, feces, stomach contents, gizzard, gizzard contents, adipose, rape, cabbage, onion, or tomato, 0.3 g of maize, 0.05 g of hair, feather, or sugar cane, 0.01 g of soil, or 1.0 mL of blood were dried for 48 h in an oven at 50 C. The dried samples were placed in pre-washed digestion vessels, followed by acid digestion using 6 mL of nitric acid (atomic absorption spectrometry grade, 60%, Kanto Chemical Corp., Tokyo, Japan) and 1 mL of hydrogen peroxide (Cica reagent, 30%, Kanto Chemical Corp.). 0.1 g of dried bone was placed in the vessels, followed by acid digestion using 6 mL of nitric acid. The digestion vessels were capped and placed onto a 10-position turntable, and subsequently underwent a ramped temperature program in a closed microwave extraction system, the Speed Wave MWS-2 microwave digestion system (Berghof, Germany). After cooling, extracted solutions were transferred into 15 mL plastic tubes and diluted to a ﬁnal volume of 10 mL with bi-distilled and de-ionized water (Milli-Q). The water samples were simply preserved with nitric acid to obtain acidic conditions. The concentrations of various elements (Pb, Cd, As, Zn, Co, Mn, and Cr) were determined using an inductively coupled plasma-mass spectrometer (ICP-MS: 7700 series, Agilent Technologies, Tokyo, Japan). Analytical quality control was performed using the DORM-3 (ﬁsh protein, National Research Council of Canada, Ottawa, Canada) and DOLT-4 (dogﬁsh liver, National Research Council of Canada) certiﬁed reference materials. Replicate analysis of these reference materials showed good recoveries (95e105%). The instrument detection limit was 0.001 mg/L. 2.3. Analysis of Pb-IRs Analyses of the 208Pb/206Pb and 207Pb/206Pb ratios were conducted using ICP-MS, according to the following procedure (Nakata et al., 2015). Detailed analytical conditions are given in supporting information of Materials and Methods as well as in the Table S2. During the analytical procedure, the following isotopes were measured: 204Pb, 206Pb, 207Pb, and 208Pb. However, only the 208 Pb/206Pb and 207Pb/206Pb ratios are discussed in this study, as they show the most signiﬁcant differences between the contaminated and natural background materials. Moreover, these ratios have been the most commonly interpreted ratios in previous research (Monna et al., 1998). The ratios of the samples were
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corrected every 10 samples using the average value of each isotope ratio obtained by measurement of SRM 981. Each sample was measured in 10 replicates. The standard error for the 208Pb/206Pb and 207Pb/206Pb measurements was generally <0.5% of RSD (relative standard deviation), except <1.0% of RSD in some of the samples which showed very low Pb concentration. We conﬁrmed that there was no Pb contamination through the analytical procedures using the regent blank measurement. 2.4. Statistical analysis All the statistical analyses were carried out using JMP 11 (SAS Institute, Cary, NC, USA) in order to evaluate signiﬁcant differences in the data. The TukeyeKramer test and Student's t-test were used to compare element accumulation levels in the biological samples from the goats and chickens, respectively. TukeyeKramer test was carried out in each species in order to compare the Pb-IRs among several biological samples from the same sampling site. To determine the relationship between Pb-B and Pb-IRs in the biological samples from the goats and chickens, spline smoothing, a method for ﬁtting a smooth curve to a set of data, was utilized (l ¼ 0.05). All of the statistical analyses were performed at the signiﬁcance level of 0.05 (p < 0.05). 3. Results 3.1. Element concentrations in biological samples from goats and chickens Concentrations of Pb, Cd, As, Zn, Co, Mn, and Cr in the blood, liver, kidney, lung, spleen, brain, muscle, heart, bone, feces, stomach contents, hair, and adipose of the three groups of goats were determined (supporting information Table S3). Levels of these elements in the blood, liver, kidney, lung, spleen, brain, gizzard, gizzard contents, muscle, heart, bone, feces, feather, and adipose of BC and FRC were also determined (supporting information Table S4). Pb levels in the liver, kidney, and lung (equivalent to z1.75, 1.98, and 0.86 mg/kg, wet wt., assuming a moisture content of 68.7, 78.7, and 79.5%, respectively) of G0 exceeded the residual limit of 0.5 mg/kg, wet wt. for offal for human consumption (Regulation Council, 2001). As for muscle, 60% of the G0 samples exceeded the 0.1 mg/kg, wet wt. maximum Pb level for human consumption (FAO, 2012). In contrast, the mean level of Cd in all of the goat samples was below the maximum level of 1,000 mg/kg, wet wt (Regulation Council, 2001). For chickens, Pb levels in the liver, kidney, lung, and brain (equivalent to z1.25, 3.44, 1.17, and 0.70 mg/kg, wet wt. assuming a moisture content of 71.9, 75.1, 77.4, and 78.9%, respectively) of the FRC exceeded the 0.5 mg/kg maximum Pb level for offal for human consumption (Regulation Council, 2001). Additionally, 20% of the FRC muscle samples exceeded the residual Pb limit of 0.1 mg/kg, wet wt. in muscle for human consumption (FAO, 2012). In contrast to goats, Cd concentrations in liver and kidney (equivalent to z1210 and 2620 mg/kg, wet wt., respectively) of the FRC exceeded the maximum level of 1000 mg/kg, wet wt. for offal for human consumption (Regulation Council, 2001). As indicated in supporting information Table S3, G0 accumulated signiﬁcantly higher concentrations of Pb in the blood, liver, kidney, lung, spleen, brain, muscle, bone, feces, stomach contents, and hair than the other groups. Signiﬁcantly higher accumulations of As and Co were also observed in G0 compared with G30 and G150. In contrast, few signiﬁcant differences in Cd, Zn, Mn, and Cr levels were found between the goat groups. As shown in supporting information Table S4, signiﬁcantly greater accumulation of Pb was observed in all of the biological samples from the FRC. In
contrast to the goat groups, the Cd level in the FRC was signiﬁcantly higher than in the BC for most of the sample types, except for blood and bone. Signiﬁcantly higher accumulations of As and Co were also observed in many tissues from the FRC with respect to the BC. 3.2. Concentrations of Pb and Cd in environmental samples Concentrations of Pb and Cd in the environmental samples are shown in supporting information Table S5. The levels of Pb and Cd in S0 exceeded the benchmark values (USEPA, 2005; Fairbrother et al., 2007). S30, S150, and the sawdust contained lower amounts of Pb and Cd than A0. The Pb and Cd levels in various vegetables from A0 exceeded or were comparable to food reference values (FAO, 2012). Maize (Zea mays) from A0 had higher levels of Pb and Cd than maize from A30 and A150. In contrast to the soil and vegetables, the levels of Pb and Cd in the water samples were lower in A0. 3.3. Pb-IRs in biological samples from goats and chickens Geographic trends in the Pb-IRs (208Pb/206Pb and 207Pb/206Pb ratios) from goats and chickens are shown in Fig. 1, and Fig. 2, respectively. Additionally, mean ± standard deviation (SD) and minimumemaximum values of the Pb-IRs from goats and chickens are shown in supporting information Tables S6 and S7. G150 exhibited large variation in the 208Pb/206Pb and 207Pb/206Pb ratios among the different tissues, ranged from 1.979 to 2.131 and 0.736 to 0.891, respectively (Fig. 1A). The Pb-IRs in the G30 and G0 tissues ranged from 2.098 to 2.208 and 2.119 to 2.166 for 208Pb/206Pb and from 0.846 to 0.890 and 0.873 to 0.886 for 207Pb/206Pb, respectively (Fig. 1B, C). As the distance to the mining site became shorter, the Pb levels in the goats increased (supporting information Table S3), and the differences in the Pb-IRs among the tissues became smaller (Fig. 1AeC). The variation in the Pb-IRs in BC was smaller than in G150, ranging from 2.086 to 2.127 and 0.850 to 0.882 for 208Pb/206Pb and 207 Pb/206Pb, respectively (Fig. 2A). Additionally, in contrast to G150, the relationship between the 208Pb/206Pb and 207Pb/206Pb ratios in BC did not indicate a linear trend. FRC, which accumulated higher levels of Pb, showed smaller variation in Pb-IRs than BC, ranging from 2.123 to 2.151 and 0.866 to 0.876 for 208Pb/206Pb and 207 Pb/206Pb, respectively (Fig. 2B). The trends in the Pb-IRs for G0 and FRC were quite similar, although the trends for G150 and G30 were different from BC, because of the latter's non-linear appearance. 3.4. Pb-IRs in environmental samples Pb-IR results for the environmental samples are shown in Fig. 3 and supporting information Table S8. Although the sample size was small, a positive relationship was observed between soil and maize Pb-IRs. Pb-IRs in both soil and maize increased as the distance to the mining site became smaller. Pb-IRs in environmental samples ranged from 2.082 to 2.465 and 0.856 to 0.885 for 208Pb/206Pb and 207 Pb/206Pb, respectively. Variation in the Pb-IRs among the environmental samples from A0 was smaller than that for G0 and FRC. 3.5. Relationship between Pb-B and Pb-IRs in various tissues from goats and chickens In the present study, both the 208Pb/206Pb and 207Pb/206Pb ratios in the liver, kidney, lung, spleen, brain, blood, muscle, heart, and bone of G30 and G150 showed large variations at 1.18 mg/dL of PbB, while the Pb-IRs in those tissues of G0 showed limited variation at 8.06 mg/dL of Pb-B (Fig. 4 and supporting information Fig. S2).
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Fig. 1. Pb-IRs (208Pb/206Pb and 207Pb/206Pb) in G150 and S150 (A), G30 and S30 (B), and G0 and S0 (C). Red diamond ¼ liver, blue circle ¼ kidney, black triangle ¼ lung, white inverted triangle ¼ spleen, blue diamond ¼ brain, asterisk ¼ blood, black square ¼ muscle, white triangle ¼ heart, red inverted triangle ¼ bone, black diamond ¼ feces, white square ¼ stomach contents, red triangle ¼ hair, blue inverted triangle ¼ adipose, S ¼ soil. (For interpretation of the references to color in this ﬁgure legend, the reader is referred to the web version of this article.)
In contrast, this trend was not observed in chickens (Fig. 5 and supporting information Fig. S3). 4. Discussion 4.1. High accumulation of metals in goats, chickens, and environmental samples High accumulations of Pb and Cd in soil (Ikenaka et al., 2010; Nakayama et al., 2011), wild rats (R. rattus) (Nakayama et al., 2011, 2013), livestock such as cattle and chickens (Yabe et al., 2011, 2013; Ikenaka et al., 2012), and children (Yabe et al., 2015) from Kabwe have been previously reported. The present study also revealed a high accumulation of Pb and Cd in the edible organs of G0 and FRC (supporting information Table S3 and S4). This study is noteworthy as the ﬁrst to reveal the extent of metal contamination in goats from Zambia. Goats are a common livestock animal in Zambia, and it was found that Pb concentrations in the liver, kidney, and lung of G0 (supporting information Table S3) exceeded the food reference value (Regulation Council, 2001). What must be heeded is that the breeding period of G0 in the mining site was less than one
month. These results should be a grave warning for farmers and consumers living in and around the mining site. There is a clear need to avoid consumption of contaminated goat offal and muscle, as well as to restrict goats from roaming and scavenging for food near mining sites. As Yabe et al. (2013) previously reported, also in the current study, high levels of Pb and Cd were also observed in FRC (supporting information Table S4), and the Pb and Cd levels in several tissues exceeded the residual limit for human consumption (Regulation Council, 2001; FAO, 2012). However, despite the accumulation of high Pb, the sampled FRC superﬁcially appeared healthy. This is in agreement with ﬁndings from previous research, in which adult hens showed tolerance to chronic Pb intoxication (Mazliah et al., 1989). Nonetheless, consumption of Pbcontaminated chicken offal and muscle poses signiﬁcant health risks to humans, especially children, who are highly susceptible to Pb toxicity (Lockitch, 1993). High accumulation of Pb and Cd was noted in S0 (supporting information Table S5), as Ikenaka et al. (2010) and Nakayama et al. (2011) previously observed. The levels of Pb and Cd in S0 exceeded benchmark values (USEPA, 2005; Fairbrother et al., 2007). Pb and
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Fig. 2. Pb-IRs (208Pb/206Pb and 207Pb/206Pb) in BC and sawdust (A), and FRC and S0 (B). Red diamond ¼ liver, blue circle ¼ kidney, black triangle ¼ lung, white inverted triangle ¼ spleen, blue diamond ¼ brain, asterisk ¼ blood, black square ¼ muscle, white triangle ¼ heart, red inverted triangle ¼ bone, black diamond ¼ feces, white diamond ¼ gizzard, white square ¼ gizzard contents, red triangle ¼ feather, blue inverted triangle ¼ adipose, S ¼ S0, D ¼ sawdust. (For interpretation of the references to color in this ﬁgure legend, the reader is referred to the web version of this article.)
The extent of contamination in the vegetables and water from Kabwe was newly revealed in this study, although the sample size was small (supporting information Table S5). Thus, vegetable consumption can be considered one of the exposure routes of Pb and Cd. In contrast, the levels of Pb and Cd in the water samples were low, even in A0. Therefore, the drinking water was deemed to pose almost no risk to human health. 4.2. Pb exposure route of goats and chickens, and potential health risk of human exposure to Pb
Fig. 3. Pb-IRs (208Pb/206Pb and 207Pb/206Pb) in the environmental samples. Red circle ¼ S150, blue circle ¼ S30, black circle ¼ S0, red square ¼ maize from A150, blue square ¼ maize from A30, black square ¼ maize from A0, black diamond ¼ rape from A0, black triangle ¼ sugar cane from A0, black inverted triangle ¼ cabbage from A0, white triangle ¼ onion from A0, white inverted triangle ¼ tomato from A0. (For interpretation of the references to color in this ﬁgure legend, the reader is referred to the web version of this article.)
Cd are known as toxic pollutants; for instance, renal dysfunction and bone loss have been reported in humans suffering from chronic Cd toxicity (J€ arup, 2003). It has been previously argued that the source of these toxic metals in the Kabwe area is Pb and Zn mining and smelting activity (Nakayama et al., 2011). The current study revealed that Pb and Cd pollution in the soil around the mining site is still ongoing. Blacksmith Institute (2013) noted that the site still poses an acute health risk that will require further work although the Zambian government has made signiﬁcant progress in dealing with the issue, particularly through a United States Dollar (USD) 26 millions remediation program funded by World Bank and Nordic Development Fund from 2003 to 2011.
Although previous studies of Kabwe have revealed signiﬁcant Pb pollution, the actual Pb exposure routes for animals and humans are still unclear. In Kabwe, leaded gasoline has been phased out since March 2008. Additionally, the Pb-IRs detected in biological samples from G0 and FRC in the current study generally differed from those in coals used worldwide in the previous study (1.98e2.12 and 0.81e0.87 for 208Pb/206Pb and 207Pb/206Pb, respectively) (Diaz-Somoano et al., 2009). Hence, leaded gasoline and coal would not be sources of Pb in Kabwe. As shown in Fig. 1C, Pb-IRs in S0 and the biological samples from G0 were quite similar. Additionally, the same trend was observed in S30 and G30 (Fig. 1B). These ﬁndings suggest that the Pb-IRs in goat tissues reﬂect those in soil, indicating that soil is one of the major sources of Pb exposure for goats around the Kabwe mining site. Another interesting point regarding the Pb-IRs is the relationship between S150 and the feces from G150. Liu et al. (2014) previously noted that feces samples are more suitable for tracing Pb sources in case where Pb is detected at low levels in the blood. The results of the current study show a similar trend in the Pb-IRs of feces samples (Fig. 1A). Incidentally, FRC appeared to show the same tendency as G0 and G30. Moreover, the trend in the Pb-IRs of BC, which were not highly contaminated by Pb, was similar to the corresponding FRC trend (Fig. 2). This result can be explained by the following hypothesis: the sawdust in the poultry house coincidentally had Pb-IRs close to those of S0, and the Pb-IRs of the BC were affected by the sawdust. However, exposure to sawdust does not result in the accumulation of large amounts of Pb. Therefore, while Pb contamination levels indeed differed, the Pb-IRs of the BC and FRC ended up being quite similar.
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Fig. 4. Pb-B (mg/dL) versus Pb-IRs (208Pb/206Pb and 207Pb/206Pb) in goat tissues. To determine the relationship between Pb-B and Pb-IRs in the biological samples from the goats, spline smoothing, a method for ﬁtting a smooth curve to a set of data, was utilized (l ¼ 0.05).
Since this is the ﬁrst study to identify Pb exposure routes in Kabwe, the next point of concern is human exposure to Pb. It is well known that Pb is absorbed into the human body mainly through the gastrointestinal and respiratory tracts (Berman, 1966; Neathery and Miller, 1975; Glorennec, 2006). Previous studies (Ikenaka et al., 2010; Nakayama et al., 2011) and the current (supporting information Table S5) study found soil Pb concentration that exceeded benchmark values (USEPA, 2005; Fairbrother et al., 2007), and supposed soil and sawdust could be sources of Pb exposure for goat and chicken, respectively, as described above. Given these considerations, the inhalation of dust or the accidental ingestion of soil, such as by having a meal with unwashed hands, could be a great risk for human exposure to Pb, especially in children. Additionally, the health risks to residents due to contamination of vegetables with heavy metals have been widely reported (Wang et al., 2005; Nabulo et al., 2006; Sharma et al., 2008). In Kabwe, the consumption of vegetables containing high amounts of Pb compared to food reference values (FAO, 2012) could contribute to human exposure to Pb. Furthermore, Farmer and Farmer (2000) previously reported that high levels of toxic metal contamination in livestock as well could be a signiﬁcant potential risk to human health. As the goats and chickens in Kabwe accumulated high levels of Pb in the current study, there is a clear need to consider the potential risk of human exposure to Pb via this route. Yabe et al. (2015) recently revealed signiﬁcant Pb accumulation in the blood of children in Kabwedall examined children under the
age of 7 years had Pb-B exceeding 5 mg/dLdand their great risk for Pb toxicity. Canﬁeld et al. (2003) reported that even a blood Pb level of less than 10 mg/dL can cause neurological abnormalities, such as manifested by a decreased intelligence quotient (IQ), in children. In response to the report, the US Centers for Disease Control and Prevention (CDC, 2012) revised the blood Pb “level of concern” from 10 to 5 mg/dL. Given the pathology and guidelines, it is clear that the residents of Kabwe, especially its children, are at great risk from Pb pollution. Therefore, further studies on the sources of Pb exposure for humans around the Kabwe mining site are imperative. 4.3. Biological fractionation and its threshold for Pb-IRs It has been well documented that the isotopic composition of Pb is not signiﬁcantly affected by physico-chemical fractionation €fer and Rosman, 2001; Veysseyre and Bollho, processes (Bollho 2001). Similarly, the high atomic mass of Pb and the slight differences in the mass of Pb isotopes support the notions that no signiﬁcant degree of fractionation takes place during various metalworking activities (Cheng and Hu, 2010), and that Pb isotopes do not fractionate measurably in biological systems (Rabinowitz and Wetherill, 1972; Carlson, 1996). Therefore, stable Pb isotopes provide an efﬁcient tool for identifying the sources and pathways of Pb pollution. The analysis of stable Pb isotopes is increasingly used in many ﬁelds of environmental research (Rabinowitz and Wetherill, 1972; Dolgopolova et al., 2006; Iglesias et al., 2010),
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Fig. 5. Pb-B (mg/dL) versus Pb-IRs (208Pb/206Pb and 207Pb/206Pb) in chicken tissues. To determine the relationship between Pb-B and Pb-IRs in the biological samples from the chickens, spline smoothing, a method for ﬁtting a smooth curve to a set of data, was utilized (l ¼ 0.05).
even recently extending to studies of wild animals such as marine mammals (Caurant et al., 2006) and birds (Finkelstein et al., 2010). However, observations of biological fractionation have been documented. Smith et al. (1996) revealed large differences in Pb isotope ratios between paired blood and bone samples from human subjects, and in a recent animal experiment, Wu et al. (2012) and Liu et al. (2014) recorded remarkable differences in Pb isotopic signature between several tissues, blood, urine, and feces of the SpragueeDawley (SD) rat. These notions suggest there might be biological fractionation of Pb isotopes in different biological samples. Despite these new insights, little is known about the fractionation of Pb isotopes in biological samples from living creatures. We understand Kabwe to be an area that has only one origin of Pb, namely a mine, a notion supported by the difference of Pb-IRs between in coals from all over the world and in biological samples from G0 and FRC, and the phasing out of leaded gasoline in the past as mentioned above. Hence, the current study can be regarded as a semi-ﬁeld study on continued exposure to Pb for goats and chickens. As shown in Fig. 1A, G150 exhibited a large variation in its Pb-IRs among different tissues. This result could support the concept of biological fractionation in natural setting. In contrast to G150, the Pb-IRs in the biological samples from G30 and G0, especially G0, were too similar to make such a distinction (Fig. 1B, C). This ﬁnding implies that high Pb accumulation in the body might affect biological fractionation, and could also impact the redistribution of Pb isotopes among tissues. The result shown in Fig. 4 and supporting information Fig. S2 revealed that the Pb-IRs in
all the goat biological samples are ﬁxed at a certain value (approximately 2.13 for 208Pb/206Pb and 0.88 for 207Pb/206Pb) when the Pb-B exceeds a speciﬁc level. This speciﬁc level of Pb-B was estimated to be approximately 5 mg/dL, and is considered the threshold for biological fractionation. These ﬁndings are quite similar to and support those from a previous study on rats (Liu et al., 2014). Contrary to goats in the current study and rats in the previous study (Liu et al., 2014), the chicken samples showed a different trend in their Pb-IRs. Although FRC exhibited similar tendencies to G0 and G30 in terms of their Pb-IRs, which showed limited variation and were close to the Pb-IRs of S0 (Fig. 2B), BC showed some small variation in the Pb-IRs of tissues, but seemed to have been primarily affected by the sawdust Pb content (Fig. 2A). These results of only minor variation in Pb-IRs lead to the assumption that chickens might have negligible biological fractionation. Another possibility is that the threshold for the disappearance of biological fractionation might be very low in chickens compared to goats and rats. While the actual reason remains unclear, the trend observed for the Pb-IRs of the chickens clearly differs from that of the mammals. This assumption could be one of the reasons why the relationship between the Pb-B and Pb-IRs in the chickens (Fig. 5 and supporting information Fig. S3) was unexpectedly quite different from that in the goats (Fig. 4 and supporting information Fig. S2). This difference can also be interpreted with the previous ﬁnding showing the tolerance of chickens to chronic Pb intoxication (Astrin et al., 1987). Further studies are clearly required to
H. Nakata et al. / Environmental Pollution 208 (2016) 395e403
unravel this interesting species difference in Pb-IR fractionation, as observed among chickens and mammals such as goats and rats.
studies should be conducted to clarify the usefulness of Pb isotope analysis for identifying Pb sources more accurately.
4.4. Reliability of identifying a Pb pollution source with stable Pb isotope analysis
In evaluating the usefulness of Pb isotope analysis for identifying a Pb pollution source, it should be noted that the Pb-IRs in various tissues from both G0 and FRC were quite similar, and close to those of the soil (Figs. 1C and 2B). In addition, the Pb-IRs in the goat tissues increased up to certain values at around 5 mg/dL of PbB, and were constant at >5 mg/dL of Pb-B (Fig. 4 and supporting information Fig. S2). These results were in accordance with the previous ﬁndings in rats (Liu et al., 2014). One of the well-known Pb toxicities is hematological toxicity; 5-aminolevulinic acid dehydratase (ALAD) in the blood is generally regarded as an indicator of Pb hematological toxicity (Astrin et al., 1987). The lowest-observed effect concentration (LOEC) for ALAD inhibition in mammals has been reported as 19, 43, 35, and 91 mg/dL for rats (Azar et al., 1972), dogs (Azar et al., 1972), rabbits (Falke and Zwennis, 1990), and calves (Lynch et al., 1976), respectively. For humans, the CDC (2012) recently revised the blood Pb “level of concern” from 10 to 5 mg/dL. Although any Pb isotopic study focusing on the accurate identiﬁcation of Pb sources must take into account the effect of Pb-B on biological fractionation (Liu et al., 2014), stable Pb isotope analysis can be still regarded as a reliable tool for identifying Pb pollution sources in the case of mammals having a Pb-B exceeding the level of concern (5 mg/dL). Moreover, the observed relationship of the PbIRs in S150 and G150 feces (Fig. 2A) implies that feces are the most reliable sample type for the identiﬁcation of Pb sources at a Pb-B lower than 5 mg/dL. On the other hand, source identiﬁcation using Pb isotope analysis cannot be regarded as being reliable in chickens. In this study, changes in Pb-IRs did not clearly relate to Pb-B in chickens (Fig. 5 and supporting information Fig. S3), although the actual reason remains uncertain. Moreover, the current study newly demonstrated that for chickens, the values of the natural background Pb-IRs could coincidentally be close to those of a pollution source (Fig. 2). Pb-IRs in both the natural and contaminated sites should be considered when identifying the pollution source.
H. N., S. M. M. N., J. Y., and M. I. designed research; H. N., S. M. M. N., J. Y., and A. L. performed research; H. N., S. M. M. N., H. M., W. S. D., and Y. I. analyzed data; and H. N., S. M. M. N., and M. I. wrote the paper. Conﬂict of interest The authors declare no conﬂict of interest. Acknowledgments This work was supported by Grants-in-Aid for Scientiﬁc Research from the Ministry of Education, Culture, Sports, Science and Technology of Japan awarded to M. Ishizuka (No. 24405004 and No. 24248056) and Y. Ikenaka (No. 26304043, 15H0282505, 15K1221305), and the foundation of JSPS Core to Core Program (AA Science Platforms) and Bilateral Joint Research Project (PG36150002 and PG36150003). We also acknowledge the ﬁnancial support by The Mitsui & Co., Ltd. Environment Fund and The Nihon Seimei Foundation. We are grateful to Mr. Takahiro Ichise (Laboratory of Toxicology, Graduate School of Veterinary Medicine, Hokkaido University) for technical support with metal concentration analyses using ICP-MS. Supporting information Supporting information (details on Materials and Methods, supporting ﬁgures and tables) are available free of charge via the Internet at http://pubs.acs.org. Appendix A. Supplementary data Supplementary data related to this article can be found at http:// dx.doi.org/10.1016/j.envpol.2015.10.006. References
5. Conclusions Large variation of Pb-IRs was observed in biological samples from goats and chickens, which had relatively low levels of contamination. The variation in the goats decreased with an increase in Pb-B. Moreover, the Pb-IRs in all goat samples were ﬁxed at certain values close to those of the soil, which can be regarded as the dominant source of Pb exposure at Pb-B >5 mg/dL. The results conﬁrmed that the biological fractionation of Pb isotopes should occur in goats, and that the threshold for triggering biological fractionation is at around 5 mg/dL of Pb-B, as suggested by an earlier study in rats. In addition, feces might be the most reliable sample type for identifying a Pb source at low Pb-B. In contrast, chickens did not show a clear relationship for Pb-IRs against Pb-B, or a fractionation threshold. Moreover, environmental Pb-IRs in the control and exposure groups of chickens appeared to be similar for reasons unrelated to pollution, and the Pb-IRs of the control group of chickens were also directly affected by environmental Pb-IRs. From these ﬁndings, chickens might not be a reliable animal for tracing Pb sources. The results also suggest that Pb-IRs in both the control and polluted area should be taken into account jointly for identiﬁcation of the Pb source. Given these, Pb isotope analysis can still be considered a helpful tool for the identiﬁcation of Pb sources. Nevertheless, further
Astrin, K.H., Bishop, D.F., Wetmur, J.G., Kaul, B., Davidow, B., Desnick, R.J., 1987. daminolevulinic acid dehydratase isozymes and lead toxicity. Ann. N. Y. Acad. Sci. 514 (1), 23e29. Azar, A., Trochimowicz, H.J., Maxﬁeld, M.E., 1972. Review of lead studies in animals carried out at Haskell Laboratory: two-year feeding study and response to hemorrhage study. Environmental Health Aspects of Lead. In: Proceedings of an International Symposium, pp. 199e210. Berman, E., 1966. The biochemistry of lead review of the body distribution and methods of lead determination. Clin. Pediatr. 5 (5), 287e291. Blacksmith Institute, 2010. World's Worst Pollution Problems Report 2010 Top Six Toxic Threats. In: Pap. Annual Reports. http://www.worstpolluted.org/2010report.html (accessed 21.07.15.). Blacksmith Institute, 2013. The Worlds Worst 2013: the Top Ten Toxic Threats Cleanup, Progress, and Ongoing Challenges Table. In: Pap. Annu. Reports, pp.1e19. www.worstpolluted.org/docs/TopTenThreats2013.pdf (accessed 21.07.15.). € fer, A., Rosman, K.L.R., 2001. Isotopic source signatures for atmospheric lead: Bollho the Northern Hemisphere. Geochim. Cosmochim. Acta 65 (11), 1727e1740. Canﬁeld, R.L., Henderson Jr., C.R., Cory-Slechta, D.A., Cox, C., Jusko, T.A., Lanphear, B.P., 2003. Intellectual impairment in children with blood lead concentrations below 10 mg per deciliter. N. Engl. J. Med. 348 (16), 1517e1526. Carlson, A.K., 1996. Lead isotope analysis of human bone for addressing cultural afﬁnity: a case study from Rocky Mountain House, Alberta. J. Archaeol. Sci. 23 (4), 557e567. pez, A., Addink, M., Caurant, F., Aubail, A., Lahaye, V., Van Canneyt, O., Rogan, E., Lo Churlaud, C., Robert, M., Bustamante, P., 2006. Lead contamination of small cetaceans in European watersethe use of stable isotopes for identifying the sources of lead exposure. Mar. Environ. Res. 62 (2), 131e148. Centers for Disease Control and Prevention, 2012. Low Level Lead Exposure Harms Children : a Renewed Call for Primary Prevention Advisory Committee on Childhood Lead Poisoning Prevention, pp. 1e65.
H. Nakata et al. / Environmental Pollution 208 (2016) 395e403 Cheng, H., Hu, Y., 2010. Lead (Pb) isotopic ﬁngerprinting and its applications in lead pollution studies in China: a review. Environ. Pollut. 158 (5), 1134e1146. Charalampides, G., Manoliadis, O., 2002. Sr and Pb isotopes as environmental indicators in environmental studies. Environ. Int. 28 (3), 147e151. Chow, T.J., Johnstone, M.S., 1965. Lead isotopes in gasoline and aerosols of Los Angeles Basin, California. Science 147 (3657), 502e503. Chow, T.J., Earl, J.L., 1972. Lead isotopes in North American coals. Science. 176 (4034), 510e511. Crosby, D.G., 1998. Inorganic Toxicants. In: Environ. Toxicol. Chem.. Oxford Univ. Press, New York, pp. 205e224. Diaz-Somoano, M., Kylander, M.E., Lopez-Anton, M.A., Suarez-Ruiz, I., MartinezTarazona, M.R., Ferrat, M., Kober, B., Weiss, D.J., 2009. Stable lead isotope compositions in selected coals from around the world and implications for present day aerosol source tracing. Environ. Sci. Technol. 43 (4), 1078e1085. Dolgopolova, A., Weiss, D.J., Seltmann, R., Kober, B., Mason, T.F.D., Coles, B., Stanley, C.J., 2006. Use of isotope ratios to assess sources of Pb and Zn dispersed in the environment during mining and ore processing within the OrlovkaeSpokoinoe mining site (Russia). Appl. Geochem. 21 (4), 563e579. Epov, V.N., Rodriguez-Gonzalez, P., Sonke, J.E., Tessier, E., Amouroux, D., Bourgoin, L.M., Donard, O.X.F., 2008. Simultaneous determination of speciesspeciﬁc isotopic composition of Hg by gas chromatography coupled to multicollector ICPMS. Anal. Chem. 80 (10), 3530e3538. k, A., Mihaljevi Ettler, V., Vane c, M., Bezdi cka, P., 2005. Contrasting lead speciation in forest and tilled soils heavily polluted by lead metallurgy. Chemosphere 58 (10), 1449e1459. Fairbrother, A., Wenstel, R., Sappington, K., Wood, W., 2007. Framework for metals risk assessment. Ecotoxicol. Environ. Saf. 68 (2), 145e227. Falke, H.E., Zwennis, W.C.M., 1990. Toxicity of lead acetate to female rabbits after chronic subcutaneous administration. 1. Biochemical and clinical effects. Arch. Toxicol. 64 (7), 522e529. Farmer, A.A., Farmer, A.M., 2000. Concentrations of cadmium, lead and zinc in livestock feed and organs around a metal production centre in eastern Kazakhstan. Sci. Total Environ. 257 (1), 53e60. Finkelstein, M.E., George, D., Scherbinski, S., Gwiazda, R., Johnson, M., Burnett, J., Brandt, J., Lawrey, S., Pessier, A.P., Clark, M., 2010. Feather lead concentrations and 207Pb/206Pb ratios reveal lead exposure history of California condors ́ (Gymnogyps californianus). Environ. Sci. Technol. 44 (7), 2639e2647. Food and Agriculture Organization of the United Nations, 2012. Joint FAO/WHO Food Standards Programme, Codex Committee on Contaminants in Foods. Sixth session. CF/6 INF/1, pp. 1e94. Gannes, L.Z., Del Rio, C.M., Koch, P., 1998. Natural abundance variations in stable isotopes and their potential uses in animal physiological ecology. Comp. Biochem. Physiol. A Mol. Integr. Physiol. 119 (3), 725e737. Glorennec, P., 2006. Analysis and reduction of the uncertainty of the assessment of children's lead exposure around an old mine. Environ. Res. 100 (2), 150e158. Gulson, B.L., Tiller, K.G., Mizon, K.J., Merry, R.H., 1981. Use of lead isotopes in soils to identify the source of lead contamination near Adelaide, South Australia. Environ. Sci. Technol. 15 (6), 691e696. €ppel, V., 2000. Lead-isotopes as tracers of pollutants in soils. Hansmann, W., Ko Chem. Geol. 171 (1), 123e144. Hudson-Edwards, K.A., Macklin, M.G., Brewer, P.A., Dennis, I.A., 2008. Assessment of Metal Mining-contaminated River Sediments in England and Wales. Environment Agency, pp. 1e64. Ikenaka, Y., Nakayama, S.M.M., Muzandu, K., Choongo, K., Teraoka, H., Mizuno, N., Ishizuka, M., 2010. Heavy metal contamination of soil and sediment in Zambia. Afr. J. Environ. Sci. Technol. 4 (11), 729e739. Ikenaka, Y., Nakayama, S.M.M., Muroya, T., Yabe, J., Konnai, S., Darwish, W.S., Muzandu, K., Choongo, K., Mainda, G., Teraoka, H., Umemura, T., Ishizuka, M., 2012. Effects of environmental lead contamination on cattle in a lead/zinc mining area: changes in cattle immune systems on exposure to lead in vivo and in vitro. Environ. Toxicol. Chem. 31 (10), 2300e2305. Iglesias, M., Sanchez, M., Queralt, I., Hidalgo, M., Margui, E., 2010. Sequential extraction combined with isotopic analysis as a tool for studying lead contamination from mining activity. Int. J. Environ. Waste Manag. 5 (1), 64e78. €rup, L., 2003. Hazards of heavy metal contamination. Br. Med. Bull. 68 (1), Ja 167e182. Kamona, A.F., Friedrich, G.H., 2007. Geology, mineralogy and stable isotope geochemistry of the Kabwe carbonate-hosted PbeZn deposit, Central Zambia. Ore. Geol. Rev. 30, 217e243. Keinonen, M., 1992. The isotopic composition of lead in man and the environment in Finland 1966-1987: isotope ratios of lead as indicators of pollutant source. Sci. Total Environ. 113 (3), 251e268. Lantzy, R.J., Mackenzie, F.T., 1979. Atmospheric trace metals: global cycles and assessment of man's impact. Geochim. Cosmochim. Acta 43, 511e525. Liu, D., Wu, J., Ouyang, L., Wang, J., 2014. Variations in lead isotopic abundances in Sprague-Dawley rat tissues: possible reason of formation. PLoS One 9 (2), e89805. Lockitch, G., 1993. Perspectives on lead toxicity. Clin. Biochem. 26, 371e381. Lynch, G.P., Jackson, E.D., Kiddy, C.A., Smith, D.F., 1976. Responses of young calves to low doses of lead. J. Dairy Sci. 59 (8), 1490e1494. Mazliah, J., Barron, S., Bental, E., Rogowski, Z., Coleman, R., Silbermann, M., 1989. The effects of long-term lead intoxication on the nervous system of the chicken. Neurosci. Lett. 101 (3), 253e257. Meyer, P.A., Brown, M.J., Falk, H., 2008. Global approach to reducing lead exposure and poisoning. Mutat. Res. 659 (1), 166e175.
Monna, F., Loizeau, J., Thomas, B.A., Gue, C., Favarger, P., 1998. Pb and Sr isotope measurements by inductively coupled plasma e mass spectrometer : efﬁcient time management for precision improvement. Spectrochim. Acta B. 53, 1317e1333. Nabulo, G., Oryem-Origa, H., Diamond, M., 2006. Assessment of lead, cadmium, and zinc contamination of roadside soils, surface ﬁlms, and vegetables in Kampala City, Uganda. Environ. Res. 101 (1), 42e52. Nakata, H., Nakayama, S.M.M., Ikenaka, Y., Mizukawa, H., Ishii, C., Yohannes, Y.B., Konnai, S., Darwish, W.S., Ishizuka, M., 2015. Metal extent in blood of livestock from Dandora dumping site, Kenya: source identiﬁcation of Pb exposure by stable isotope analysis. Environ. Pollut. 205, 8e15. Nakayama, S.M.M., Ikenaka, Y., Hamada, K., Muzandu, K., Choongo, K., Teraoka, H., Mizuno, N., Ishizuka, M., 2011. Metal and metalloid contamination in roadside soil and wild rats around a Pb-Zn mine in Kabwe, Zambia. Environ. Pollut. 159 (1), 175e181. Nakayama, S.M.M., Ikenaka, Y., Hamada, K., Muzandu, K., Choongo, K., Yabe, J., Umemura, T., Ishizuka, M., 2013. Accumulation and biological effects of metals in wild rats in mining areas of Zambia. Environ. Monit. Assess. 185 (6), 4907e4918. Neathery, M.W., Miller, W.J., 1975. Metabolism and toxicity of cadmium, mercury, and lead in animals: a review. J. Dairy Sci. 58 (12), 1767e1781. Petit, D., Mennessier, J.P., Lamberts, L., 1984. Stable lead isotopes in pond sediments as tracer of past and present atmospheric lead pollution in Belgium. Atmos. Environ. 18 (6), 1189e1193. Rabinowitz, M.B., Wetherill, G.W., 1972. Identifying sources of lead contamination by stable isotope techniques. Environ. Sci. Technol. 6 (8), 705e709. Razo, I., Carrizales, L., Castro, J., Díaz-Barriga, F., Monroy, M., 2004. Arsenic and heavy metal pollution of soil, water and sediments in a semi-arid climate mining area in Mexico. Water Air Soil. Pollut. 152, 129e152. Regulation Council, 2001. No 2375/2001 of 29 November 2001 amending Commission Regulation (EC) No 466/2001 setting maximum levels for certain contaminants in foodstuffs. Off. J. Eur. Communities L. 321 (1), 6e12. Sangster, D.F., Outridge, P.M., Davis, W.J., 2000. Stable lead isotope characteristics of lead ore deposits of environmental signiﬁcance. Environ. Rev. 8 (2), 115e147. Scheuhammer, A.M., Templeton, D.M., 1998. Use of stable isotope ratios to distinguish sources of lead exposure in wild birds. Ecotoxicology 7 (1), 37e42. Sharma, R.K., Agrawal, M., Marshall, F.M., 2008. Heavy metal (Cu, Zn, Cd and Pb) contamination of vegetables in urban India: a case study in Varanasi. Environ. Pollut. 154 (2), 254e263. Shirahata, H., Elias, R.W., Patterson, C.C., Koide, M., 1980. Chronological variations in concentrations and isotopic compositions of anthropogenic atmospheric lead in sediments of a remote subalpine pond. Geochim. Cosmochim. Acta 44 (2), 149e162. Smith, D.R., Flegal, A.R., Niemeyer, S., Estes, J.A., 1990. Stable lead isotopes evidence anthropogenic contamination in Alaskan sea otters. Environ. Sci. Technol. 24 (10), 1517e1521. Smith, D.R., Osterloh, J.D., Flegal, A.R., 1996. Use of endogenous, stable lead isotopes to determine release of lead from the skeleton. Environ. Health Perspect. 104 (1), 60. n, H., Malinovsky, D., Engstro €m, E., Rodushkin, I., Baxter, D.C., Stenberg, A., Andre 2004. Isotopic variations of Zn in biological materials. Anal. Chem. 76 (14), 3971e3978. Stokes, P.E., Okamoto, M., Lieberman, K.W., Alexander, G., Triana, E., 1982. Stable isotopes of lithium: in vivo differential distribution between plasma and cerebrospinal ﬂuid. Biol. Psychiatry 17 (4), 413e421. Tieszen, L.L., Boutton, T.W., Tesdahl, K.G., Slade, N.A., 1983. Fractionation and turnover of stable carbon isotopes in animal tissues: implications for d13C analysis of diet. Oecologia 57, 32e37. United States Environmental Protection Agency, 2005. Guidance for Developing Ecological Soil Screening Levels, pp. 1e85. Veysseyre, A.M., Bollho, A.F., 2001. Tracing the origin of pollution in French alpine snow and aerosols using lead isotopic ratios. Environ. Sci. Technol. 35 (22), 4463e4469. Walczyk, T., von Blanckenburg, F., 2002. Natural iron isotope variations in human blood. Science 295 (5562), 2065e2066. Wang, X., Sato, T., Xing, B., Tao, S., 2005. Health risks of heavy metals to the general public in Tianjin, China via consumption of vegetables and ﬁsh. Sci. Total Environ. 350, 28e37. Wu, J., Liu, D., Xie, Q., Wang, J., 2012. Biological fractionation of lead isotopes in Sprague-Dawley rats lead poisoned via the respiratory tract. PLoS One. 7 (12), e52462. Yabe, J., Nakayama, S.M.M., Ikenaka, Y., Muzandu, K., Choongo, K., Mainda, G., Kabeta, M., Ishizuka, M., Umemura, T., 2013. Metal distribution in tissues of free-range chickens near a lead-zinc mine in Kabwe, Zambia. Environ. Toxicol. Chem. 32 (1), 189e192. Yabe, J., Nakayama, S.M.M., Ikenaka, Y., Muzandu, K., Ishizuka, M., Umemura, T., 2011. Uptake of lead, cadmium, and other metals in the liver and kidneys of cattle near a lead-zinc mine in Kabwe, Zambia. Environ. Toxicol. Chem. 30 (8), 1892e1897. Yabe, J., Ishizuka, M., Umemura, T., 2010. Current levels of heavy metal pollution in Africa. J. Vet. Med. Sci. 72 (10), 1257e1263. Yabe, J., Nakayama, S.M.M., Ikenaka, Y., Yohannes, Y.B., Bortey-Sam, N., Oroszlany, B., Muzandu, K., Choongo, K., Kabalo, A.N., Ntapisha, J., Mweene, A., Umemura, T., Ishizuka, M., 2015. Lead poisoning in children from townships in the vicinity of a lead-zinc mine in Kabwe, Zambia. Chemosphere. 119, 941e947.