4. BIOAVAILABILITY AND TOXICITY OF CADMIUM
4.2. Risk to Animals and Humans
The tolerable intake for Cd as proposed by FAO/WHO (1972) is 400–500 àg week−1 person−1 or 57–71 àg day−1 person−1 weighing 70 kg. Despite the difference in the amount of dietary Cd intake, however, rice is the lead- ing source of Cd common to many populations in Asia, e.g. approximately ranging from 20% of total dietary Cd intake in the Philippines (Zhang et al., 1998), to 30–40% in Japan (Ikeda et al., 1999; Watanabe et al., 2000a), many parts in mainland China (Zhang et al., 1997), Taiwan (Ikeda et al., 1996), and Thailand (Zhang et al., 1999), and with a much higher level (53%) in Malaysia (Moon et al., 1996). It may be of interest to stress that wheat (Triticum aestivum L.) and other cereals are additional sources of dietary Cd when people consume such cereals more than rice, as observed in north- ern China (Watanabe et al., 1998, 2000b). It is, however, worth noting that while rice is the major component of food among the Asian community, many of them (especially females and children) weigh <70 kg. Thus FAO/
WHO tolerable limit for Cd and indeed other potentially toxic substances may not be appropriate for Asian communities.
In comparison, dietary Cd intake levels in the 1990s in Europe have been made available in recent years. The levels reported (mostly geomet- ric mean or medians for adult people, unless otherwise specified) include 6.3 àg week−1 person−1 for 1.5–5.3-year-old children in a remote island in the North Sea, Germany (Schrey et al., 2000), approximately 7 àg day−1
in Duisburg, Germany (Wilhelm et al., 1995), 9–22 àg day−1 in Sweden ( Jọrup et al., 1998), 11.1 àg day−1 for students in Spain (Barbera et al., 1993), 14 àg day−1 in the UK (Ysart et al., 1999), 17.5 àg day−1 in Cro- atia ( Sapunar-Postruznik et al., 1996), 19–27 àg day−1 in Belgium (Van Cauwenbergh et al., 2000), and 28 àg day−1 in the Ruhr district, Germany (Wilhelm et al., 2002). The value reported for Canadians (Dabeka and McKenzie, 1995), 13 àg day−1, was also close to these values observed in Europe. The levels as a whole appear to be close to or slightly higher than the levels in Asia (5–15 àg day−1; Ikeda et al., 2000) excluding Japan where the exposure level is much higher (i.e. ∼30 àg day−1).
In particular, background exposure of general populations in Japan to Cd has been relatively high (Watanabe et al., 2000a). According to Watanabe et al. (1996), there is a significant difference in geometric mean Cd content in rice samples (1546) collected in 17 areas in the world, but it was the high- est in Japan (0.0557 mg kg−1) and Colombia (0.13320 mg kg−1) (Table 4.11).
Analysis of data from food duplicate surveys in the mid 1990s on some 600 women throughout the country coupled with air pollution data disclosed that foods were the almost exclusive source of Cd exposure for the general population, whereas the exposures through the respiration route was essen- tially negligible; rice (paddy rice), the staple cereal for most Japanese, was in fact the leading source of Cd intake via daily foods, accounting for 30–40%
in the 1980s and 1990s (Ikeda et al., 1999; Watanabe et al., 2000a), although a substantial decrease in total dietary intake occurred in the mid 1990s as compared with the levels 10–15 years ago (Watanabe et al., 1996, 2000a).
The Japanese people used to take in approximately 8.2 and 1.8 àg Cd day−1 from rice and wheat, respectively (Shimbo et al., 2001) (Table 4.12). It is worth noting that the dietary route is responsible for 99% of
Table 4.11 Cadmium Contents in Rice Grain from Various Areas (Watanabe et al., 1996) Areas Mean Cd (mg kg−1) Areas Mean Cd (mg kg−1)
Australia 0.00267 Vietnam 0.01850
China 0.01554 Canada 0.02902
Taiwan 0.03955 Colombia 0.13320
Indonesia 0.02177 Finland 0.02580
Japan 0.05570 France 0.01741
Korea 0.01570 Italy 0.03392
Thailand 0.01570 South Africa 0.02582
Malaysia 0.02774 Spain 0.00085
Philippines 0.02014 USA 0.00743
total Cd intake in Japan because Cd concentration in general atmospheric air is low compared to food (Ikeda et al., 2000). The dietary Cd intake by middle-aged Japanese women in the late 1990s was 25.5 àg day−1 ( Watanabe et al., 2000a) (Table 4.13). This amount of Cd intake in Japan appeared to be 2–3 times higher than the levels among other women populations in east and south-east Asia (Ikeda et al., 2000), possibly because Cd contents in the rice are lower there, even though in these places people depend even more heavily on rice as a major source of energy for daily life.
Cadmium reaches the food chain also through uptake by grazing ani- mals. For example, in New Zealand, most of the Cd accumulated in grazing animals is derived mainly from the pasture intake (Loganathan et al., 2008).
Bramley (1990) has estimated that annually ∼55 and 275 mg Cd is ingested per sheep and cow, respectively, through the intake of herbage. Work in New Zealand has indicated that cattle and sheep may ingest 1–10% and
>30%, respectively, of their dry matter intake in the form of soil. In areas where soils contain high Cd, for example, as a result of sewage sludge and fertilizer application, soil ingestion is expected to play a significant role in Cd uptake by farm animals (Loganathan et al., 2003; Wilkinson et al., 2001).
Roberts et al. (1994) have shown that even though sheep ingest 36–46 kg soil per year, this only contributes 5–8% of the total Cd intake for the lax and hard grazed flocks, respectively.
Ruminants do not have a homeostatic control mechanism for regulat- ing Cd absorption or excretion which is affected by the level of dietary Cd content (Lee et al., 1996; Loganathan et al., 2008). Although intes- tinal uptake of Cd has been estimated to account for over 90% of the
Table 4.12 Intake of Cadmium by Japanese via Rice and Wheat in 1998–2000
Item
Rice Wheat
Average Ranges Average Ranges
Daily cereal consumption (g day−1)
165 158–178 91 67–104
Cadmium content (mg kg−1)*
0.0497 0.0430–0.0701 0.0193 0.0170–0.0212
Daily Cd intake
(àg day−1) 8.2 7.409–12.506 1.754 1.297–1.993
*Cited from Ministry of Health and Welfare (2000) Revised from Shimbo et al. (2001)
total Cd absorbed, about 80–90% and 0.05% of the total ingested Cd is excreted in the faeces and urine, respectively. Most dietary Cd is bound to metallothionein and is absorbed intact into the circulation, and, in animals, kidney and liver Cd accounts for 50–70% of the total Cd with kidney having a higher Cd concentration than the liver. Other organs such as the pancreas, spleen, heart, brain, and testis, together with muscle and fat accumulate small amounts of Cd. In blood, Cd is associated with albumin-like protein, which is transported to the kidney where it is fil- tered through the glomerulus and reabsorbed by the proximal tubules (Friberg et al., 1985).
In animals, the effects of prolonged exposure to abnormal levels of Cd include testicular necrosis, placenta destruction, abortion, teratogenic malformations, renal damage, osteomalacia, immunosuppression, pulmo- nary edema, and emphysema. In humans, severe Cd exposure can result in emphysema, bronchitis, ulceration of nasal mucosa, renal dysfunction, liver necrosis and anemia, hypertension, skeletal deformities, prostrate and lung cancer, and teratogenesis (Chowdhury and Chandra, 1987).
Table 4.13 Dietary Exposure to Cadmium (Cd), and Cd Concentration in Blood and Urine among Nonsmoking Adult Women in the Cities in Asia
Location
Cadmium in
References Food
(àg day−1) Blood (àg L−1) Urine (mg kg−1 Creatinine) Japan (A group of 27 sites)
1977–1981 37.5 3.47 – Watanabe et al.
(2000a)
1991–1997 25.5 1.90 4.39 Watanabe et al.
(2000a)
Bangkok 7.1 0.41 1.40 Zhang et al.
(1999) China (A group
of 4 sites) 9.9 1.07 2.30 Zhang et al.
(1997) Korea (A group
of 4 sites) 21.2 1.39 2.26 Moon et al.
(1998)
Kuala Lumpur 9.0 0.74 1.51 Moon et al.
(1996)
Manila 14.2 0.47 1.21 Zhang et al.
(1998)
Tainan 9.7 0.83 1.59 Ikeda et al.
(1996)
Revised from Watanabe et al. (2000a)
Different countries have set guidelines on the maximum permissible levels for Cd in various meat products. In a survey of cattle, pigs, and sheep in New Zealand, the mean Cd concentration in kidney cortex, liver, and muscle was <0.4, 0.1, and 0.05 mg Cd kg−1 wet weight, respectively (Solly et al., 1981). These values are well below both the concentration measured in other countries and the maximum permissible concentration stipulated by many countries. However, Roberts et al. (1994) indicated that between 1988 and 1991, some 22–28% of sheep and 14–20% of cattle had kidney Cd contents greater than the permissible level of 1 mg Cd kg−1. In general, older animals had higher kidney Cd contents as these animals had longer exposure to Cd in their environment and, hence, greater opportunity to consume and retain Cd (Petterson et al., 1991). However, the rate of Cd increase decreases with age (Lee et al., 1996); for example, at 6 months of age, sheep retain about 0.1% of the Cd intake in the kidneys, with this value decreasing to 0.04% or less with increasing age (Loganathan et al., 2008).
Feral deer and sheep isolated from human intervention also showed age- related retention of Cd in kidney tissue as a result of exposure to naturally occurring Cd in the environment.
Cadmium in cow and goat milk (0.002–0.006 àg kg−1) (Rayment, 1988) and sheep and cattle muscle (0.01–0.03 mg kg−1 wet weight) were reported to be very low. Cadmium measured in carcass meat in New Zea- land (0.02–0.03 mg kg−1 wet weight) (Roberts and Longhurst, 2002) and Australia (median, 0.01 mg kg−1 wet weight) (Langlands et al., 1988) were also significantly lower than the permissible level of 0.05 mg kg−1 fresh tissue (kidneys, liver, muscle) in these countries (Loganathan et al., 2008).
Because the accumulation of Cd in kidneys was high, whereas the amounts in milk and meat were low, Loganathan et al. (2008) suggested that the graz- ing animal can be viewed as a “filter” capable of ensuring that the amounts of Cd entering the food chain are low. They also conveyed that this was similar to the case, in which Cd-rich bran and roots were removed from processed grains and edible parts of plants, respectively, and then less Cd entered the food chain (Sauerbeck, 1992).
The average annual consumption of red meat per New Zealander (70 kg adult) is estimated to be 80 kg. Based on a median Cd content of 0.01 mg kg−1 wet weight, the annual intake of Cd is expected to be 0.8 mg, which is equivalent to a daily intake of 2.19 àg Cd. This is considerably less than the maximum safe level of 70 àg Cd day−1 (Ledgard et al., 2010;
Loganathan et al., 2008) from all sources in the diet. Although offal, such as liver and kidney, contains more Cd than the other meat, the small quantity
of offal consumed by the New Zealand population is unlikely to contribute significantly to the total Cd intake. Also, as most meat-producing animals are slaughtered at a young age, this ensures that Cd concentrations of their edible offal products are within permissible limits (Loganathan et al., 2008).
However, in sections of population and in pets where offal forms a larger portion of the diet, Cd intake from this source could be a major concern (Loganathan et al., 1999).
The “population critical concentration” of Cd at which 10% of the human population would exhibit signs of renal impairment on the kidney is about 160 mg Cd kg−1 wet weight (Friberg, 1984). From the above calcu- lations, it can be shown that it is unlikely that Cd input from meat exceeds the critical concentration in the life time of a New Zealander. However, Loganathan et al. (2008) proposed that computer-based models are required to identify farming systems that present a high risk of Cd concentrations in edible offal exceeding the permissible levels and livestock at risk of chronic toxicity. They recommended that a decision support model of this kind may be useful in developing management strategies capable of reducing Cd accumulation in animals. They stated that although preliminary empirical models have been developed for Cd accumulation in sheep grazing on New Zealand pastures (Loganathan et al., 1999), further development of these models is required for their wider applicability.