Food Safety Issues for Mineral and Organic Fertilizers

CHAPTER TWO

Rufus L. Chaney

Contents 1. Introduction 52 2. Concentrations of Trace Elements in Fertilizers and Soil Amendments 53 3. Risk Assessment Pathways 57 4. Natural Controls on Trace Element Transfer from Soil to Plants 58 5. Element Risks 59 5.1. Selenium 59 5.2. Molybdenum 60 5.3. Cobalt 61 5.4. Cadmium 62 5.5. Lead 83 6. Trace Elements in Other Common Soil Amendments 84 6.1. Manure 84 6.2. Limestone 85 6.3. Steel production fume waste 86 6.4. Gypsum 86 7. Long-Term Reactions of Trace Elements in Soils 87 8. Other Elements in Fertilizers and Soil Amendments of Possible 89 Concern 8.1. Fluoride 89 8.2. Radionuclides 92 8.3. Arsenic 94 9. Monitoring and Control of Trace Elements in Mineral Fertilizers 97 10. Need for Regulatory Enforcement on Composition of Soil 98 Amendments References 99

Senior research Agronomist, USDA-Agricultural Research Service, Environmental Management and Byproducts Utilization Lab, BARC-West, Beltsville, MD, USA

Abstract Fertilizers and other soil amendments are required to maintain soil fertility, but some may be naturally rich in trace elements, or contaminated. As part of the overall consideration of using fertilizers and soil amendments, one should consider the levels of trace elements present in relation to soil, plant, and food-chain processes (precipitation, adsorption, chelation) which promote or alleviate trace element risks. These natural processes limit plant accumulation of nearly all elements to levels which would not cause harm to humans, livestock, wildlife, or soil organisms. Soils geologically rich or contaminated with Mo can harm ruminants, while those rich in Se may harm all plant consumers; Mo or Se should be applied only when needed. Manures from swine and poultry may be rich in Cu, Zn and/or As from feed additives. Crops except rice accumulate little As from soils, so soil ingestion is the basis for soil As risk except for rice. Pb risk is also through soil ingestion rather than plant uptake. Cd is accumulated by rice to levels which caused human disease (renal tubular dysfunction) where rice soils were contaminated by industrial discharges. Risk from Cd in rice is strongly affected by the high bioavailability of rice Cd. Consumption of similar amounts of Cd have not caused harm from other foods. Because phosphate fertilizers may contain high levels of Cd, and use of high Cd superphosphate in Australia caused signiﬁcant increase in wheat and potato Cd levels, risk from long-term accumulation of phosphate fertilizer Cd (and other sources) must be controlled. Different control schemes are discussed.

1. Introduction

Cases of excessive food-chain transfer of soil or applied trace elements indicate that all fertilizers and soil amendments require evaluation. Some products may be contaminated even when nearly all similar products are quite low in trace elements. Thus, control systems are required to assure compliance with quality standards to protect purchasers and food safety. On the other hand, soil and crop properties have prevented transfer of excessive trace elements in nearly all situations, so details are needed to understand why some elements and products have caused adverse effects in some cases while most do not. Historically, veterinary toxicology has shown that certain elements in soil or amendments can poison livestock if the element is applied in excess, or occurs naturally at high levels in the soil. In particular, Se and Mo have been known to harm livestock grazing mineralized alkaline soils for longer than 100 years (McDowell, 2003; Underwood, 1977). More recently, CdeZn mine waste contamination of rice soils in Asia caused excessive accumulation of Cd but not Zn in rice grain, such that subsistence farm families suffered Cd health effects (Kobayashi, 1978; Nogawa et al., 2004). Because Cd was long known to occur in phosphate fertilizers Food Safety Issues for Mineral and Organic Fertilizers 53

(e.g., Williams and David, 1973), these findings in Asia caused concern about food-chain transfer of fertilizer-applied Cd. But research has helped explain why significant Cd disease has only been found in subsistence rice consumers with industrially-contaminated paddy soils (Chaney, 2010). Great concern has been expressed about some fertilizer products which contain high levels of Cd, Pb, and As, although nearly all As and Pb concern is about garden fertilizer products which may be ingested by children rather than garden crop uptake of As and Pb. More recently, excessive ingestion of sulfate from well water, soil amendments, or feed additives has caused a sulfur toxicity disease in ruminants (Gould, 1998). These examples of identified problems from excessive food-chain transfer of trace elements in agriculture raised many questions about how the diseases occurred, and how better management of fertilizers and soil amendments could prevent future problems. Much research has been conducted on these issues for decades, giving us a better understanding of the soileplanteanimal relationships of trace elements. In addition, improved science supported risk assessment by government agencies that sorted out which elements and products need controls, and which should be handled by management practices. On the other hand, excessive concern based on poor science could lead to unnecessarily restrictive limits on trace elements in fertilizers, and raise the costs of these products and of foods to consumers. This paper does not consider biosolids use in agriculture. I have reviewed the topic of biosolids and biosolids composts in several recent publications (Chaney and Ryan, 1993; Chaney and Ryan, 1994; Chaney et al., 2001; Basta et al., 2005). It should be recognized that intense research on biosolids has supported the development of understanding of science about soil and amendment trace element risk assessment (US-EPA, 1995).

2. Concentrations of Trace Elements in Fertilizers and Soil Amendments

Speciﬁc fertilizer and soil amendment products (phosphates; byproduct Zn fertilizers) normally contain signiﬁcant levels of elements of potential concern (Cd, As, Pb, F, etc.), while others (N and K fertilizers) normally contain only very low levels of these elements (Table 1). Table 2 shows that phosphates from different sources contain a wide range of Cd concentrations depending on the rock phosphate source and processing technology (McLaughlin et al., 1996; Chen et al., 2007; Grant et al., 1999). Among soil amendments other than pesticides, arsenic is 54

Table 1 Arsenic and cadmium concentrations in commercial fertilizers in California

As Cd Mean Median Range Mean Median Range L Fertilizer Type mg kg 1 P Fertilizers 11 13 0e21 89 132 0e163 Rock Phosphates 7.4 7.5 1.4e13 33 0.5 0e130 Multi-nutrient Fertilizers 13 8.0 0.15e155 37 19 0e200 Biosolids Fertilizers 2.4 1.7 0e10 2.8 1.0 0e15 Zn Micronutrient 30 0.5 0e280 149 23 0e495 Sulfur Micronutrient 5.8 1.0 0.1e19 0.43 0 0e3.0 Iron Micronutrient 1544 118 0.3e4950 71 21 0e334 ZneFeeMn Blends 48 e 24e71 248 e 95e400

(CDFA, 1998). Chaney L. Rufus Table 2 Composition and limits for Cd in P fertilizers in several countries Fertilizers Organic and Mineral for Issues Safety Food

Cd in 45% Cd:P Cd:P2O5 P2O5-Product L mg Cd kg 1 L1 L1 Country Limit mg Cd kg P P2O5 mg Cd kg Limits for Fertilizer-Cd Washington 0.0889 kg Cd ha1 yr1 2040 889 400 1 USA-Oregon 7.5 mg Cd (%P2O5) 774 338 152 1 USA-California 4 mg Cd >(%P2O5) 412 180 81 1 AAPFCO 10 mg Cd (%P2O5) 1030 450 202 Australia 300 mg Cd (kg P)1 300 131 59 Canada 0.0889 kg Cd ha1 yr1 2040 889 400 Japan 340 148 67 1 Austria 75 mg Cd (kg P2O5) 275 120 54 1 Belgium 90 mg Cd (kg P2O5) 206 90 40.5 Denmark 110 48.0 21.6 Netherlands 40 17.5 7.9 1 Finland 21.5 mg Cd (kg P2O5) 49 21.5 9.7 1 Sweden 43 mg Cd (kg P2O5) 100 43.7 19.7 EU Proposal (2001) 1 20 mg Cd kg P2O5 45.8 20 9.0 1 40 mg Cd kg P2O5 91.6 40 18 1 60 mg Cd kg P2O5 137.0 60 27 (continued) 55 56 Table 2 (continued)

Cd in 45% Cd:P Cd:P2O5 P2O5-Product L mg Cd kg 1 L1 L1 Country Limit mg Cd kg P P2O5 mg Cd kg Levels present in example products Kola 1.2 0.523 0.24 USA-FL 46 20 9 Sweden 118 51 23 (TSP) USA-NC 336 147 66 Nauru 641 280 126 High-Western-US 1170 511 230

(Grant et al., 1999; data for Swedish traditional phosphates from Andersson and Hahlin, 1981). uu .Chaney L. Rufus Food Safety Issues for Mineral and Organic Fertilizers 57 only above background soil concentrations in a few phosphate products and some livestock manures. But phosphate products produced from most ores contain Cd above the levels in soils except for electric furnace phosphate products. The concentrations are expressed in different ways because different groups conducting risk assessment focus on Cd:P ratio, Cd:P2O5 ratio, or concentration of Cd in the actual product. There is general acceptance that limits for Cd or other elements in P-fertilizers should be expressed per unit P or P2O5. Chien and Menon (1994) compared rock phosphate, acidulated rock phosphate, and normal phosphate fertilizers and stressed that one should also consider Cd on the basis of Cd per unit phytoavailable P; P phytoavailability varies widely among rock phosphates as both P and Cd have varied solubility, but the Cd in rock phosphate eventually becomes labile according to several studies (Kuo et al., 2007).

3. Risk Assessment Pathways

In order to evaluate whether trace elements in speciﬁcfertilizersor soil amendments might comprise risk, one must evaluate the potential for the element to cause adverse effects to organisms be they humans, livestock, wildlife, or soil microbes. Over the last few decades, pathways for assessment of potential risks have become established during the risk assessments for land application of numerous materials. The pathways to humans were always clear: the general agricultural pathway focused on grains; the garden foods pathway focused on fruits and vegetables grown in home gardens; the livestock pathways to humans; water, air, and soil- derived dust. Livestock and wildlife health are assessed both for plants grown on amended soils, and for ingestion of the amended soil or the amendment itself adhering on plant leaves. Wildlife have different pathways than humans and livestock, and bioaccumulation may play a greater role in wildlife food-chains than for humans; because earthworms accumulate Cd within their bodies, earthworm consuming wildlife become a highly-exposed species for soil Cd enrichment. The pathways have been well described and used in risk assessments in the US and the EU (US-EPA, 1995). Careful evaluation of all the pathways for trace element transfers and risks allows appropriate assessment of which levels of contaminants are required to protect food safety and the environment. But appropriate assessment requires evaluation of interactions which limit element transfer and risk, not just the worst possible cases which are often stressed by those seeking more strict controls. 58 Rufus L. Chaney

4. Natural Controls on Trace Element Transfer from Soil to Plants

Natural properties of elements, soils, plants, and animals greatly limit the potential for risk from nearly all elements in a process I have called the “Soil-Plant Barrier” (Chaney, 1983; Langmuir et al., 2005; McLaughlin et al., 1999). Most elements are so insoluble in soils or in roots that they do not reach the edible shoots of plants at levels which could cause even chronic toxicity to plants or consumers (e.g., Cr3þ, Hg, F, Pb, Ag, Ba, Zr, V, rare earths, etc.). Other elements may be taken up and readily translocated to edible plant tissues, but phytotoxicity limits the maximum concentration in edible plant tissues with visible injury to levels which are chronically tolerated by livestock, wildlife, and humans (Zn, Cu, Ni, Mn, F, As, B). Reduced yield or death of the plant limits exposure of consumers to these elements in soil. Livestock are usually the most exposed and susceptible plant consumers because they may consume 100% of their diet as plants which are experiencing significant yield reduction due to phytotoxicity. And ruminants are sensitive to some elements (Mo, Se, S) for which non-ruminants are much more tolerant. A few elements are readily absorbed and translocated to plant shoots without causing phytotoxicity. These are the elements which are known to potentially cause food-chain toxicity (Mo, Se, Cd, and possibly Co). These elements are considered in some detail herein. Another process allows exposure of animals to elements in soils, the direct ingestion of soil or soil amendment. Grazing livestock are known to consume appreciable amounts of soil either as splash on leaves, or over-grazed pastures, or as amendments adhering to plant leaves (e.g., Chaney and Lloyd, 1979). And young children ingest housedust (partly soil which has been carried into homes by parents, siblings or pets) by hand-to-mouth play. Risk assessment research has found that young children ingest an average of 24 4 (SD) mg soil d1 with the 95th percentile ingestion rate of 91 16.6 mg d1 (Stanek et al., 2001). Even if elements such as Pb, As, or F are not accumulated by plants to dangerous levels because of the chemistry of soils and plants, the soil element may be sufficiently bioavailable in ingested soil to cause adverse effects to livestock, wildlife, or even humans. Elements for which soil or soil amendment ingestion is the principal source of potential risk to humans or livestock include: Pb, As, F, Fe2þ,Cr6þ and Hg. Many elements occur in fertilizers and soil amendments at measurable levels, but in most products the elements are at low concentrations and cannot cause food-chain contamination issues even after centuries of use. But elements in some materials require controls to prevent the identified Food Safety Issues for Mineral and Organic Fertilizers 59 potential problems. All products must be protected from deliberate contamination. Understanding why and how some elements need to be controlled in fertilizers and soil amendments helps with understanding how humans, livestock and wildlife are protected against this potential risk by control systems or natural processes.

5. Element Risks

5.1. Selenium The Mo and Se poisoning of livestock are complicated, but illustrate soil- plant-animal processes which allow risk. First, Se is well known as a natural toxin in geologically mineralized soils in several US states. On the one hand, most US soils are deficient in Se for human and livestock health (Kubota et al., 1987). Increasing interest in improving Se levels in human diets has resulted in tests of Se fertilization (e.g., Bañuelos, 2006; Broadley et al., 2 2010). Because selenate ðSeO4 Þ is an anion, and it is strongly adsorbed on soil hydrous Fe oxide at low soil pH, plant uptake of Se is very limited in acidic soils, but can reach high levels in alkaline soils if Se is present. The other half of Se risk through food-chain transfer is that plants are very tolerant of Se, such that plants tolerate higher levels than do the animals that consume the crops. A typical maximum chronic dietary level for livestock is 2 mg kg1 dry weight (NRC, 1980), while forage plants can accumulate 100 mg kg1 without harm to the plant. And there are even hyperaccumulators of Se which reach 10,000 mg Se kg1 (Chaney et al., 2010). Selenate is usually absorbed by the sulfate transporter of roots, so if sulfate is simultaneously high in soils, selenate accumulation is strongly reduced (Stroud et al., 2010), except for the hyperaccumulators (Bell et al., 1992). Concern was raised about Se in land-applied fly ash from coal-fired power plants. If Se is present in the coal, much Se enters the fly ash collected during exhaust filtration. When plants were grown on fly ash landfills, the plants contained enough Se to poison wildlife (Arthur et al., 1992). But when similar fly ash was applied at fertilizer or limestone replacement rates, the Se only raised plant Se to levels sufficient for livestock for a few years (Gutenmann et al., 1979). Se is more dangerous in aquatic than terrestrial ecosystems because the Se can be bioaccumulated in algae and biomagnified in lower animals because it can be built into proteins as selenomethionine and selenocys- teine. These problems were discovered at the Kesterson Reservoir in the Central Valley of California where Se-enriched drainage waters in evaporation ponds caused severe adverse Se effects on wildlife. The Se was leached from irrigated Se-mineralized soils which did not cause 60 Rufus L. Chaney accumulation of excessive Se levels in forages. But when evaporation raised concentrations, and aquatic organisms accumulated and then biomagnified Se, poisoning of wildlife followed (Ohlendorf et al., 1986). These observations caused wide concern about excessive Se applications, while large numbers of humans and livestock remained with inadequate Se for best health (Kubota and Allaway, 1972; Kubota et al., 1987; Broadley et al., 2010). Thus Se fertilizer applications are limited to those required to satisfy food quality goals, and to limit potential runoff, leaching, or excessive accumulation in plants or aquatic ecosystems. Livestock diets are routinely supplemented with Se, and most of that Se ends up in manure applied on cropland. But no problems have been reported from Se in manure of normally Se-supplemented livestock manures. However, in an unusual case in China, ash from burning of Se- mineralized coal was applied on cropland; maize (Zea mays L.) accumulation of Se poisoned both livestock and humans (Yang et al., 1983). Observing the livestock health problems should have allowed diagnosis of the Se problem, but the location was remote and human Se toxicity disease occurred.

5.2. Molybdenum Risks from Mo have traditionally occurred where poorly-drained Mo- mineralized soils were alkaline because this element is present as the 2 molybdate anion ðMoO4 Þ (Kubota et al., 1987; O'Connor et al., 2001). Molybdate is bound strongly by Fe oxides in acidic soils, but is more soluble and readily accumulated by many plants in alkaline soils to concentrations which can poison ruminant livestock but have not been found to comprise risk to non-ruminants. In the rumen, sulfate is reduced to sulfide and reacts with Mo to form thiomolybdates. Thiomolybdates form non-bioavailable complexes with Cu in the rumen (or in the body) and induce Cu deficiency when plant tissues contain over about 10 mg Mo kg1 dry weight. Mo toxicity is complex because the concentration and form, and hence bioavailability of Cu in plants varies, and other soil factors can inhibit Cu absorption by ruminants (O’Connor et al., 2001). Extremely high sulfate in ruminant diets can also induce Cu deficiency and other adverse effects (see below). So risk assessment for Mo is interrelated with the management of Cu and S levels of diet and water, and levels of soil ingestion which can also inhibit Cu absorption (Suttle et al., 1984). Locations with high geogenic Mo and alkaline soils often cannot be used for ruminant livestock production because of the potential for forages or grains to accumulate excessive Mo. Cu supplementation or injection can readily counteract excessive Mo ingestion when it is diagnosed. Food Safety Issues for Mineral and Organic Fertilizers 61

One example of a soil-plant-food chain Mo problem involved smelter emission of Mo in PA where some nearby soils were strongly acidic. For many years the low uptake of Mo from acidic soils prevented any adverse effects. But when the soils were limed, Mo uptake was greatly increased and cattle were harmed (Hornick et al., 1977). Neighboring farms which had used limestone application for many years to improve soil fertility had no Mo toxicity because the higher pH promoted leaching of Mo from the root zone. Plants are highly tolerant of Mo. When acidic soils require Mo fertilization to support growth of plant species with higher Mo requirements (e.g., cauliflower, legumes), it may be useful to add limestone to increase the phytoavailability of Mo instead, or in addition to Mo fertilizers so that future excessive Mo for ruminants will not occur upon liming. Legumes commonly accumulate higher levels of Mo than do grasses, so the specific crops being grown on mineralized or fertilized soils must be considered in preventing problems. A comprehensive consideration of allowable Mo levels in land-applied biosolids discusses all factors known to interact with potential Mo excess in ruminants (O’Connor et al., 2001). If biosolids containing no higher than 40 mg Mo kg1 are annually applied at N-fertilizer rates to fertilize a mixture of forage species, no adverse effects are predicted based on numerous field tests of Mo phytoavailability from biosolids. If soils are alkaline which promotes Mo uptake, the same soil pH promotes leaching from the root zone and alleviates Mo risk over time (Phillips and Meyer, 1993). One biosolids with extreme Mo contamination from an industrial source would have caused severe Mo toxicity in ruminant livestock if it had been used for forage production (Pierzynski and Jacobs, 1986).

5.3. Cobalt The case of soil Co risk to ruminants is theoretical; no adverse effects have been observed in agricultural practice. Rumen microbes require Co for normal metabolism. But when fed with diets more than 10 mg Co kg1 DW, young cattle can be poisoned by Co (NRC, 1980; Ely et al., 1948). This concentration is lower than the concentration required to cause visual symptoms of Co phytotoxicity in most plant species (25e50 mg kg1), so the “Soil-Plant Barrier” does not prevent Co poisoning of ruminants (Chaney, 1983). However, few soils or soil amendments contain high levels of Co, and excessive Co in forage diets has not been reported to date. Legumes also require Co for N ﬁxation, and some soils require Co fertilization for full yield of some legumes. So this theoretical risk from excessive Co requires attention for evaluation of Co-rich soil amendments and Co fertilizer rates. Natural soils rich in Co include serpentine soils with 100e400 mg Co kg1, but the same soils contain about 10-fold higher Ni levels which inhibit Co uptake, and can 62 Rufus L. Chaney cause Ni phytotoxicity before plants can attain even 10 mg Co (kg DW)1. This interaction is illustrated in Malik et al. (2000). In other cases, some trace elements (Cu, Co) enriched P-fertilizers have adhered to leaves of forages until washed off by rainfall; grazing livestock consumed excessive trace elements and were harmed (Underwood, 1977). Co and Cu fertilizer applications for grazed pastures require attention to prevent this possible direct ingestion of the fertilizer Co and Cu. These were reported only in Australia where extensive trace element deﬁcient pasture land has to be fertilized to support livestock production, but the livestock were not removed to other pastures during the fertilizer application.

5.4. Cadmium The last of the potentially food-chain toxic trace elements is cadmium. Although Cd can cause phytotoxicity, plants can accumulate high enough Cd concentration to comprise risk to humans before phytotoxicity is evident. Because healthy plants can accumulate enough Cd to harm humans over 50 years of chronic dietary exposure, we must limit Cd in soils and soil amendments. Adverse effects of soil Cd on human health were first observed in Japan where farm families consumed home-grown rice for decades from paddies where Cd- and Zn-rich mine wastes were deposited from upstream sources (e.g., Kobayashi, 1978). Rice grain accumulated over 1 mg Cd kg1 (e.g., Takijima and Katsumi, 1973; Fukushima et al., 1973). The disease was called Itai-itai (ouch-ouch) because of the pain from repeated bone fractures of some in the population (about 220 of more than 16,000 with adverse effects on proximal kidney tubule function at Toyoma, Japan) (Aoshima et al., 1995; Nogawa et al., 2004). Subsequently, similar Cd kidney disease has been found in many populations consuming home- grown rice from contaminated paddy soils in Japan and China, and one in Thailand (Cai et al., 1990; Simmons et al., 2003), and bone disease was also found at one location in China ( Jin et al., 2004). Based on many studies, one can expect Cd disease in populations exposed to rice grown on soils near ZnePb mines and smelters, and any upstream Cd industry, because water-borne Cd is retained effectively in paddy soils (e.g., Simmons et al., 2005, 2009). Based on the concerns raised by the observed Cd disease in Japan, many scientists studied potential soil and food Cd risks. However, food-chain risks from Cd are very complex, and many readers have not taken this complexity into account. It has become evident that Cd in different foods has quite different bioavailability. For example, Cd in shellfish has been consumed in high amounts without causing unusual Cd absorption by humans and without harm (Sharma et al., 1983; Sirot et al., 2008; Copes Food Safety Issues for Mineral and Organic Fertilizers 63 et al., 2008). Further, individuals with home gardens on soils highly contaminated by Zn mine or smelter wastes have not been found to suffer Cd disease (Strehlow and Barltrop, 1988; Ewers et al., 1985; Sarasua et al., 1995; de Burbure et al., 2003). Zn which accompanies Cd in these contaminated soil both inhibits Cd uptake and Cd bioavailability, but also can cause severe phytotoxicity at low leaf Cd compared to Cd contamination alone, limiting Cd in lettuce (Lactuca sativa L.) leaves to lower than 5mgkg1 DW (Chaney and Ryan, 1994). In order to better understand the risks of Cd in foods, controlled feeding tests with humans and laboratory animals have been conducted. Many toxicologists have fed very high Cd dietary concentrations so they could see adverse effects quickly, but high dietary Cd cannot mimic the potential chronic absorption of Cd in real foods. Fox et al. (1979) discuss in detail why levels of Cd found in foods grown on Cd-enriched soils are needed for valid research on food Cd risks. Much of the available feeding test literature is simply irrelevant because the diet Cd may have been higher than total diet Zn rather than the usual 1% or less of dietary Zn in nearly all foods. More recently it was found that Cd is predominantly absorbed by the ferrous transporter in the duodenum (DMT1) (Bannon et al., 2003; Kim et al., 2007; Park et al., 2002), helping to explain why even mild iron deficiency stress caused higher Cd absorption by humans (Flanagan et al., 1978). Increased dietary Zn can also inhibit Cd absorption by competition for the transporter (e.g., Jacobs et al., 1978). Zn grown into leafy vegetables significantly reduced Cd absorption from the crop as well (McKenna et al., 1992), and Swiss chard grown on control vs. biosolids amended soils with up to 5-times higher crop Cd and Zn did not cause any increase in kidney or liver Cd in Guinea pigs fed a high fraction of chard in their diet for a long period (Chaney et al., 1978). Although Cd is considerably higher in bran and whole wheat than in white flour, Cd absorption was not higher from the bran and whole wheat (Moberg et al., 1987; Wing, 1993). These studies suggested that the fiber and phytate reduced Cd absorption from bran and whole wheat, while their other studies showed that added Fe reduced Cd retention significantly from all of the bread diets (Wing et al., 1992). With this evidence, and the understanding that polished rice is deficient in Fe, Zn, and Ca for human health, Reeves and Chaney (2008) undertook experiments to clarify how the supply of Fe, Zn, and Ca in a crop food may affect absorption of Cd from that food, using levels of Cd present in the crop normally consumed by humans. Rats fed polished rice diets with marginal levels of Fe, Zn, and Ca (not low enough to reduce the rate of gain, but low enough to induce up-regulation of transport proteins in the intestine) absorbed 10-fold more Cd than rats fed adequate levels according to standard nutritional diets for lab animals (Reeves and Chaney, 2002). But because sunflower kernels provide more Zn, Fe, 64 Rufus L. Chaney phytate, and fiber than polished rice, absorption from sunflower kernels was significantly less than from rice in diets, and the marginal diets had less effect on Cd absorption in sunflower than rice diets (Reeves and Chaney, 2001). They then did a study to follow the fate of Cd from a single rice meal over time in rats fed marginal vs. adequate ZneFeeCa diets, finding that the marginal diets caused about 10-fold higher absorption into the duodenum, and that the Cd remained in the duodenum up to about 10 days post feeding in the marginal animals which allowed the 10-fold higher Cd release to blood and liver and subsequent slow transfer to kidney (Reeves and Chaney, 2004). With this evidence of the large difference in how rats handled Cd vs. Zn from rice diets, Reeves et al. (2005) tested whether metallothionein (MT) was involved in Cd absorption. Toxicologists have long suggested that absorbed Cd was bound by MT, a low molecular weight cysteine-rich protein found in intestine, kidney, liver, etc., and which is known to bind stored Cd in the kidney. Using mice null for the MT gene, fed marginal and adequate ZneFeeCa diets, they showed that the absence of the MT gene had no effect on Cd absorption by the animals, and no interaction with the adequacy of ZneFeeCa in the diet on Cd absorption. Thus, the long claimed important role of MT in Cd metabolism was shown to be irrelevant to food Cd risks. MT is important in preventing acute toxicity of dietary or injected Cd, but unrelated to chronic dietary Cd exposure. Strong direct evidence whether the specific food or diet markedly affects whether Cd in that food actually comprises risk has been obtained from human volunteer feeding studies. It is well known that Cd intake from smoking shows up in the blood within 24 hrs (e.g., Ellis et al., 1979). Blood indicates recent absorption of Cd from diets, so sampling individuals consuming diets with low vs. high levels of Cd intakes from specific foods can test whether such foods have bioavailable Cd. This is clearly shown in the report of Sharma et al. (1983) in which smoking increased blood Cd much more than consuming large amounts of oyster Cd. In tests with young non-smoking women in Sweden, shellfish consumers ingested 3-times higher daily Cd on average than omnivores who did not consume shellfish; but they had no higher blood Cd (Vahter et al., 1996). Shellfish are excellent sources of bioavailable Fe and Zn that reduce dietary Cd bioavailability. Similarly, volunteers consuming twice as high dietary Cd with daily sunflower kernels than peanuts had no higher blood Cd during a year-long feeding test (Reeves et al., 2001). Reeves and Chaney (2008) reviewed the evidence that different foods have quite different bioavailability of Cd in the food, indicating that bioavailability should be considered in Cd risk assessment (not presently considered in JECFA or CODEX deliberations). Other aspects of dietary Cd risk assessment are confusing to many readers, but the science is clear that diet-induced Cd disease is very unusual Food Safety Issues for Mineral and Organic Fertilizers 65 except for rice subsistence farmers over 50 years old. Chaney and Ryan (1994) summarized Cd exposures from soils in different populations vs. measures of Cd effects noting that the normal distribution of urinary indicator proteins extends well beyond the levels used by European researchers as the cut-off between normal and adverse effect levels. The excretion of b2-microglobulin (b2MG) by humans normally increases with age, such that “normal” b2MG levels in urine reach over 1 1000 mgg creatinine for 50e60 year olds, and over 1500 mg b2MG g1 creatinine for 60e70 year olds (Kowal and Zirkes, 1983; Chaney and Ryan, 1994). But European researchers commonly define disease 1 starting at 300 mg b2MG g creatinine. And Japanese experience clearly shows that persons with actual renal tubular dysfunction due to excessive 1 Cd ingestion often excrete urine containing 100,000 mg b2MG g creatinine (Aoshima, 1987; Ikeda et al., 2003; Tohyama et al., 1981). Further, the cut-off for normal (“not yet injured by Cd”) is for the upper 95% Confidence Interval, which presumes that those above the cut-off are affected while there is no evidence of Cd effect until much higher b2-MG is attained. In Japan, China and Thailand contaminated areas, individuals with frank diet Cd-induced renal tubular dysfunction comprise over 50% of the highly-exposed subsistence rice consuming population over 50 years of age (Cai et al., 1990; Honda et al., 2010; Horiguchi et al., 2010; Nogawa et al., 2004). While in the European studies which claim to show adverse effects of Cd on the population at low Cd intakes have at most a few percent of the population (Åkesson et al., 2005; Alfven et al., 2000; Järup and Alfvén. 2004; Olsson et al., 2002; Suwazono et al., 2006), likely the fraction of the normal population above the 95% Confidence Interval. And northern Europeans have among of the lowest food Cd intake levels, population kidney Cd concentration (Olsson et al., 2002), and kidney cortex Cd levels (Friis et al., 1998) among developed nations. Claims from the European Food Safety Agency (EFSA, 2009, 2010) that the allowable daily intake of Cd should be reduced to 2.5 mg (kg body weight)1 wk1 were rejected by JECFA (2010) which slightly reduced their long standing recommendation of 7 mg Cd (kg BW)1 wk1 to 25 mg Cd (kg BW)1 month1 (30 days) (stressing the chronic aspect of Cd risk requiring high consumption of foods containing higher than normal levels of bioavailable Cd for decades before adverse effects could occur). The EFSA Opinion is contraindicated by human study results in Japan where middle-aged non-smoking urban women with much higher average Cd intake than their counterparts in Europe have no evidence of Cd disease even though the high end of their normal urinary Cd excretion is 3 mgg1 creatinine (Ikeda et al., 2006). Another key misunderstanding of Cd risk is whether the higher consumption of Cd per unit body weight by children is fully accounted 66 Rufus L. Chaney for in the JECFA evaluation. Some scientists stress concern about the risk from daily Cd intake by children compared to the JECFA Potentially Tolerable Weekly Intake (PTWI), but such concern is inappropriate for the lifetime chronic risk from dietary Cd. Children normally consume as much as double the Cd daily intake in terms of mg Cd (kg BW)1 d1 compared with adults consuming the same foods because they are growing and require more food intake per unit body weight than adults. But the model for 50 year chronic exposure required to reach the maximum acceptable kidney cortex Cd concentration is based on the 50 year food consumption pattern of normal foods and food Cd levels of a normal Swedish population (Friberg et al., 1985; Kjellström and Nordberg, 1978). Children's higher Cd consumption rate is fully included in present JECFA guidance based on chronic adult diet Cd intake. The evidence about actual rice Cd-induced renal tubular dysfunction, and the rare bone disease in those with severe Cd-induced renal tubular dysfunction, coupled with the evidence about low Cd bioavailability from shellfish and sunflower kernels compared with that of rice, indicate that consumption of excessive bioavailable Cd in western diets is much more difficult to achieve than in rice diets of subsistence farmers. Appro- priate consideration of allowable safe chronic daily Cd consumption is required in order to select safe Cd levels in crops, safe soil pH management, and safe levels of Cd in agricultural amendments. CODEX established a limit for Cd in rice grain at 0.4 mg Cd kg1 fresh weight (CODEX, 2006), but maintained the limit for wheat at 0.2 mg Cd kg1. This is irrational in consideration of the large volume of data showing that rice comprises far higher potential for Cd risk than wheat. And in the case of durum wheat, much of the whole grain Cd is removed with the bran during milling to make durum flour. But the durum wheat is shipped as whole grain to prevent spoilage, so the whole grain Cd limit unnecessarily limits importation of durum wheat which will produce low Cd foods. Further, while many persons consume locally grown rice grain where the whole crop could have been grown on contaminated soils, few if any individuals consume only home-grown wheat products. Taking into account the high bioavailability of Cd in rice grain, the limit for Cd in wheat should be higher than the limit for rice.

5.4.1. Many factors affect Cd accumulation by crops Identification of limits for Cd in fertilizers is remarkably complex because so many different aspects of agriculture can influence Cd accumulation and bioavailability. It is generally agreed that the first adverse effect of excessive lifetime Cd ingestion is renal tubular dysfunction of humans. Absorption of crop Cd is strongly affected by human nutritional status, the bioavailability of crop Cd (especially Fe, Zn, and Ca status of the exposed individual and the diet), and the presence of fiber and phytate in the diet. Humans may Food Safety Issues for Mineral and Organic Fertilizers 67 ingest crops which contain high levels of Cd from production on Cd- enriched soils, and the potential for risk is related strongly to the years of high Cd crop ingestion, and the fraction of diet grown on high Cd soils. Tobacco is another very significant source of human kidney Cd with smoking of normal cigarettes doubling kidney Cd at age 50 compared to non-smokers (Elinder et al., 1976). Natural soils vary widely in total Cd, from below 0.1 to over 2 mg kg1 in normal agricultural soils, and over 100 mg kg1 in rare highly Cd mineralized soils. Soils may be contaminated from very different sources of Cd including: aerosol emissions of industrial sources; Cd in irrigation water contaminated by industrial sources; manure; biosolids; composts; fertilizers (especially P-fertilizers); and other soil amendments. The Cd:Zn ratio of the source can vary from very low (commercial purified Zn), through Zn ores, to P-fertilizers and Cd-only sources such as Cd-pigment and Cd-plastic stabilizer wastes. Cd in soils has very wide variation in phytoavailability due to soil, crop, and environmental/management factors. The crop and soil factors which most strongly influence crop concentrations of Cd are 1) crop species and cultivar (Grant et al.,2008); 2) soil pH; 3) soil and fertilizer Zn; 4) P-fertilizer rate and Cd concentration; 5) chloride in soil or irrigation water (McLaughlin et al.,1999); 6) use of chloride containing fertilizers; 7) rate and form of N-fertilizer; 8) soil drainage; 9) preceding legume or Cd accumulating crop species (Oliver et al.,1993); 10) tillage; and 11) Zn deficiency of crop/soil (Oliver et al.,1994). When so many factors can influence Cd accumulation in crops, and the bioavailability of Cd in crops, it becomes difficult to demonstrate a relationship between Cd concentration in a P-fertilizer and risk from that Cd (Grant and Sheppard, 2008; McLaughlin et al., 1996; Mench, 1998). Only a few papers have examined the complex multiple interactions possible between tillage, fertilizer placement, Cd concentration in the applied fertilizer, form and amount of N-fertilizer and multiple years (e.g., Gao et al.,2010), and such studies show the difficulty in interpretation of Cd phytoavailability research. The major factors (soil pH, NH4eN rate, chloride, crop, cultivar, P-rate, tillage, N-placement, etc.) can be observed in many studies, but the interactions can overwhelm the major factors when one looks at the details. All the interactions make development of recommendations for farmers quite difficult except to avoid or minimize Cd accumulation in cropland, especially rice and tobacco cropland. Another topic which should be kept in mind in evaluation of the exist- ing literature is that Cd plant uptake results of greenhouse pot experiments are not closely related to results of the same plant and soil in the field. In a comparison of greenhouse pots, outdoor microcosms and field plots, deVries and Tiller (1978) evaluated Cd and other element uptake by lettuce and onion from a soil treated with a biosolids. Uptake was 68 Rufus L. Chaney

4-times higher in greenhouse grown lettuce than field grown. This is believed to result from restriction of roots to the treated soil in pots, but not in the field or microcosms, the higher fertilizer density in pots, and the water use patterns in a greenhouse vs. outdoors. Many flaws in pot studies have been noted in Cd research; perhaps the most important common error is application of Cd without Zn, and use of high added rates of Cd that would never be allowed in the environment. Study of Cd addition at rates which induce biosynthesis of phytochelatin have been shown to be irrelevant to environmental exposures to the usual Cd þ Zn combination in the real world (McKenna and Chaney, 1995). Many misunderstandings of the effect of soil factors, food-chain factors, and human factors in risk from soil or fertilizer Cd to humans cause contro- versy about fertilizer Cd limits. The following discussion attempts to describe these factors and clarify how foods are protected from excessive Cd in nearly all circumstances. For particular crops, fertilizers with lower Cd concentrations are clearly indicated (tobacco, subsistence rice). But for other crops, extreme low Cd:P2O5 limits are not supported by research on food-chain transfer of Cd from applied fertilizers. Soil pH is usually found to be a highly significant factor in crop Cd accumulation, and because of this effect, any changes in soil pH (especially in the rhizosphere) due to form and rate of N-fertilizer may significantly affect Cd levels in plants. Literally hundreds of papers have reported the effect of decreasing pH raising crop Cd, and the effect of especially NH4eN increasing crop Cd (e.g., Grant et al., 1999; McLaughlin et al., 1996; Chaney, 2010). An interesting field test of the effect of N-fertilizer form and rate, and the Cd concentration in an applied P-fertilizer was conducted with wheat in the field in Sweden (Wångstrand et al., 2007). As expected (NH4)2SO4 caused greater Cd uptake than NH4NO3, which caused greater Cd uptake than Ca(NO3)2 (which causes the rhizosphere pH to rise). But even increased rate of Ca(NO3)2 caused increase in wheat grain Cd concentration compared with no added N-fertilizer. That increase may be due to effects of N on root growth, or increasing the Cd sink in the grain, but it is seen in many evaluations of grain Cd and grain N. Soil redox is especially important in Cd accumulation in rice grain. From the first studies of rice and Cd, it became apparent that little Cd was accumulated in rice shoots until the flooded paddy was drained (e.g., Takijima and Katsumi, 1973). Late rains reduced grain Cd. Careful studies showed that while the soil was flooded, Cd solubility and phytoavailability was very low. This appears to result both from the rise in pH and the formation of CdS in the flooded soil (Khaokaew et al., 2011). In recent research to characterize use of rice for phytoextraction of soil Cd, Murakami et al. (2009) reported that before draining the paddy in July, shoots contained less than 2 mg Cd kg1 DW, while at Food Safety Issues for Mineral and Organic Fertilizers 69 harvest, the shoots contained 70 mg Cd kg1 DW. Further, Simmons et al. (2008) showed that in order to predict Cd uptake by rice from extractable Cd in soil, it was important to sample and immediately analyze the moist field soil so that the soil redox and pH during grain filling would be compared to grain Cd. And although grain and shoot Cd can rise by more than 10-fold after draining the field, and leaf Zn can be increased, grain Zn is hardly changed after draining the field. The failure to bring Zn along with Cd into the grain is a key part of why rice can cause so much higher Cd risk than other crops which do not exclude Zn under soil conditions where Cd is accumulated. For example, in the Simmons et al. (2003) report on rice accumulation of Cd and Zn from mine waste contaminated paddies in Thailand, the brown rice grain grown in a high Cd and Zn-contaminated field (229 mg Cd and 6420 mg Zn kg1,pH 7.9) contained up to 2.64 mg Cd kg1, but only about 22.2 mg Zn kg1; but Chinese cabbage grown in the same field in another season accumulated 5.8 mg Cd and 643 mg Zn kg1 DW (Simmons, Angle and Chaney, unpublished).

5.4.2. Cd in phosphate fertilizers Nearly all P-fertilizers naturally contain higher Cd than soils. Some of the earliest data showing higher crop Cd on P-fertilized soils came from Australia where the unusual soil conditions and high Cd rock phosphate source used to produce superphosphate for Australia promoted Cd uptake by crops by the time it was identified (Williams and David, 1973). High Cd superphosphate produced from high Cd phosphate rocks from Christmas and Nauru Islands was applied across Australia for decades. Some Australian soils promote Cd uptake because of acidity, and others because of chloride (see below). Williams and David (1973) found significant Cd accumulation in tillage depth soils (well correlated with P accumulation in the soils), and higher Cd in crops grown on the fertilized soils. After other research was completed, it became evident that not all P-fertilizers or soils comprised equal risk. Over the decades, many well- conducted field experiments in various nations have evaluated the value of different phosphate fertilizer products. More recently, many of these sites were examined for plant uptake of Cd in relation to fertilizer P application rate and fertilizer Cd or Cd:P concentrations. In these studies, when the P-fertilizer contained low levels of Cd, no increase in accumulation of Cd in soils or plants were observed even after many years of cropping (Mortvedt, 1984, 1987; Mortvedt et al., 1981; Rothbaum et al., 1979; Mishima et al., 2004; Richards et al., 1988; Isermann, 1982; Sauerbeck, 1982; Singh and Myhr, 1998; Smilde and van Luit, 1983), while when high Cd concentration phosphate fertilizers were used, soil Cd increased and crop Cd may have been increased over time, but still far below crop Cd levels needed to cause excessive human exposure over decades. 70 Rufus L. Chaney

In some other studies of this question, soil from highly fertilized fields was used in pot tests of Cd uptake. When soil accumulation of Cd was appreciable, plant uptake of Cd was strongly increased by the historic P and Cd application (Mulla et al., 1980). And when high Cd phosphates were applied to light-textured strongly acidic soils in pot studies, lettuce or other crops accumulated increased amounts of Cd compared to very low Cd P-fertilizers (Reuss et al., 1978; Singh, 1990). In other field experiments of short duration, even when the applied phosphate source had relatively high Cd concentration, the applied P and Cd in the different products (63 vs. 379 mg Cd (kg P)1 on average) hadlittleornoeffectonCduptakebypotato(McLaughlin et al., 1995); but the locations where the tests were conducted had large differences in Cd accumulation in potato tubers. The present application of phosphates with varied Cd levels showed low effect on Cd in several crops while NH4eN fertilizer rate more strongly affected Cd accumulation (Grant et al., 2010; Singh, 1990; Mortvedt et al., 1981). Cd accumulation was more related to the presence of high levels of chloride from irrigation water than soil Cd or even pH (McLaughlin et al., 1999). In these studies, application of KCl as the K fertilizer did not significantly increase Cd uptake into potato tubers compared to use of K2SO4, but perhaps the KCl was overwhelmed by the chloride in irrigation water. In another field test, using KCl caused asignificant 30% increase in potato tuber Cd compared to K2SO4 (Sparrow et al., 1993, 1994). Perhaps because the high Cd:Zn ratio of P-fertilizers caused the Cd accumulated in Australian soils to have high Cd:Zn ratio, addition of high rates of Zn fertilizer (up to 100 kg Zn ha1 tested) significantly reduced Cd in potatoes (McLaughlin et al., 1995). Other field studies have shown the effect of nitrogen rate and especially the NH4eN rate on Cd uptake by many crops (Grant and Sheppard, 2008; Grant et al., 1999; 2002; McLaughlin et al., 1996). N rate usually has a significant effect even when the Cd concentration in the applied P- fertilizer has no effect on plant Cd. Even increasing P rate caused an increase in plant uptake of Cd regardless of Cd level in the P-fertilizer (Grant et al., 2000). Further, preceding legume or Cd-accumulator crops (sunflower, flax, etc.) caused a significant increase in wheat Cd in several studies (Oliver et al., 1993; Grant et al., 2010). And tillage practices may also influence crop Cd concentrations (Oliver et al., 1993). In another study, the previous crop sunflower shoots were removed from the field which reduced Cd accumulation by wheat (Khoshgoftarmanesh and Chaney, 2007). These many interactions make it difficult to predict Cd accumulation in crops based on fertilizers or other agronomic factors despite the clear addition of Cd in some high Cd concentration P-fertilizers over time (McLaughlin et al., 2000). Food Safety Issues for Mineral and Organic Fertilizers 71

While many assume that use of commercial phosphate fertilizers must cause higher Cd accumulation in crops than would organic farming methods, several field comparisons have been conducted. Jorhem and Slanina (2000) compared wheat in controlled field experiments and rye, carrots, and potatoes grown on commercial farms using either organic or conventional methods. There were no significant differences in Cd levels in rye, carrot, or potato over time, while wheat comparisons were inconsistent. They concluded that organic farming, at least in the short term, does not necessarily result in reduced levels of Cd in foods of vegetable origin. A similar outcome was reported by Harcz et al. (2007).

5.4.3. Cd in by-product zn fertilizers In the 1970s, some commercial P-fertilizers contained more than 400 mg Cd kg1. Some fertilizer manufacturing by-products with more than 1000 mg Cd kg1 were marketed in the US (CDFA, 1998). After a newspaper article in Seattle, WA, warned about the potential for contamination of fertilizers manufactured from hazardous industrial byproducts (Wilson, 2001), wide concern was raised. The product in question was manufactured from a hazardous fume waste of electric furnace steel production which contained substantial levels of Zn, Pb, Cd, As, and dioxins (US-EPA, 2002). A risk assessment conducted by US-EPA determined that limits were needed for such industrial byproduct derived fertilizer products, and they specified that the Zn must be separated from the byproduct to become equivalent to normal purified Zn (US-EPA, 2002). The regulation requires that Zn fertilizers manufactured from recycled hazardous materials (such as steel mill fume waste) contain no more than 1.4 mg Cd per 1% Zn in the product. For a typical ZnSO4,H2O product with 36.5% Zn, the Cd limit would be 1 50.7 mg kg ; or for ZnSO4,7H2O with 22.7% Zn, the Cd limit would be 31.8 mg kg1. The ratio of Cd:Zn thus has to be below 0.00014 which greatly limits Cd uptake by plants. Application of the unregulated products would have raised soil concentration of the contaminants. But because the Cd:Zn ratio was much lower than 1:100, excessive Cd accumulation in crops was not observed even with high cumulative application rates of these products (Mortvedt, 1985; Westfall et al., 2005). Zn in such Zn-fertilizer products inhibits uptake of Cd in the product. Increasing rates of product application over 5 repeated applications actually reduced Cd levels in lettuce shoots in a repeated application pot study by Kuo et al. (2004). The Zn product tested contained 469 mg Cd and 178,000 mg Zn kg1, with Cd:Zn ratio of 0.00263 g g1. But the application of such industrially-contaminated products on cropland was not acceptable due to the high levels of As, Pb and dioxins, and purification of the Zn from the recycled material was required if for no other reason than the dioxin present. 72 Rufus L. Chaney

Utilization of ground waste rubber has been tested as a Zn fertilizer. Tire rubber contains 1e2% Zn, and is generated in large quantities. Inter- estingly, rubber uses puriﬁed Zn such that the Zn in rubber is as low in Cd as required for byproduct Zn fertilizers by US-EPA (2002). Taheri et al. (2011) showed that ground rubber was an effective Zn fertilizer on Zn deﬁcient soils and that adding ground rubber as a Zn fertilizer could reduce Cd in wheat grain.

5.4.4. Limits on Cd in phosphate products US States started to adopt limits for trace elements in fertilizer products, and the American Association of Plant Food Control Officials (AAPFCO) adopted recommended limits (Table 2). Other nations promulgated and the EU considered limits for Cd in P-fertilizer products with a schedule to reduce the 1 allowable concentrations from 60 to 40 to 20 mg Cd (kg P2O5) (equal to 133, 89, and 44 mg kg1) product in triplesuperphosphate products containing 45 % P2O5 (EU, 1988) [see also the explanation of how these limits were developed in Cupit et al. (2002) and de Meeûs et al. (2002)]. But in 2011, the EU has not yet adopted regulations on Cd in P- fertilizers; EU expert groups reviewing the proposal suggested that the 20 mg Cd kg1 P limit was more restrictive than needed to protect food safety (CSTEE, 2002). Some perspective on Cd and other trace elements in fertilizers is needed. First, historically, some fertilizer products have contained high levels of trace elements (Table 1)(Chenet al., 2007; Mortvedt, 1996; McLaughlin et al., 1996). A few fertilizers were contaminated by industrial sources, but some phosphate ores were naturally rich in Cd from their marine sources (Grant et al., 1999; McLaughlin et al., 1996, 1999; Nziguheba and Smolders, 2008) (Table 2). Others (e.g., Kola) were low or very low because of their ore source. Some were essentially devoid of Cd (electric furnace manufactured products) due to the production methods. But electric furnace phosphates are considerably more expensive than products made using sulfuric acid and they are no longer used in most commercial fertilizer products. So it is important to look at the evidence about potential risk from fertilizer Cd in human foods. As noted above, rice diets promoted Cd disease because Cd in rice had higher bioavailability than Cd in other foods. Further, although nearly all geogenic source Cd has high levels of accompanying Zn (Cd:Zn ratio lower than 0.010), rice grain was not increased in Zn when grown on the contaminated paddy soils (Fukushima et al., 1973; Simmons et al., 2003). Other grains and vegetables always have a large increase in Zn when Cd is increased by common Cd sources to cropland. P-fertilizers have less Zn per unit Cd than Zn ores (usual Cd:Zn ratio in phosphates is about 0.16 vs.<0.010 in Zn ores) and most sources of Cd contamination Food Safety Issues for Mineral and Organic Fertilizers 73 except NieCd battery manufacturing wastes or Cd-pigment production or disposal wastes (Chaney, 2010). Another crop with unusually strong Cd accumulation is tobacco (Bell et al., 1988), and the Cd in tobacco enters the lungs of smokers and is well absorbed (Elinder et al., 1976). Tobacco is a stronger source of Cd in human kidney than is food excepting rice grown on contaminated soils. So Cd enrichment of rice and tobacco cropland require special attention. Tobacco grown on mine waste contaminated soils contributed significantly to the Cd health effects at one location in China (Cai et al., 1990), and in Thailand. Production of tobacco on Cd-enriched soils is contraindicated, and it has been recommended that P-fertilizers for use in tobacco production should be low in Cd (Lugon-Moulin et al., 2006). As noted above, many studies have been reported for Cd accumulation by crops grown on long-term phosphate fertilized soils. Because some locations received applications of Cd-rich P-fertilizer products for decades, some papers reported significant increase in crop Cd due to P-fertilizer applications (e.g., Williams and David, 1973). But in many other studies, even very long-term field experiments, no significant increase was found in crop Cd despite small increases in soil Cd (e.g., Mortvedt, 1987; Jung et al., 1979). In other cases, manure applications applied more Cd to field plots than did fertilizers and aerosol emissions, but the manure applications caused reduced levels of Cd in crops (Jones and Johnston, 1989). Looking at all common sources of Cd for cropland, Nicholson et al. (2003) found that deposition of aerosol metals exceed P-fertilizers and biosolids as Cd source in the UK soil budget (Table 3). And livestock manures ranged widely in Zn, Cu and Cd concentrations (Nicholson et al., 1999)(Table 4). In addition, different farm types use different soil amendments, hence giving different potential Cd accumulation. For example, farms which concentrate on animal production purchase off-farm feedstuffs and feed supplements and apply their manure locally (e.g., Bengtsson et al., 2003),

Table 3 Zn, Cd and Cu accumulation and Cd:Zn of accumulated metals in soils of England and Wales from different sources

Zn Cd Cu Cd:Zn L L Sources mg yr 1 gg 1 Aerosol Deposition 2457 21 631 0.0085 Livestock Manure 1858 4.2 643 0.0022 Biosolids 385 1.6 271 0.0042 Phosphate Fertilizers 213 10.0 30 0.047

(Nicholson et al., 2003). 74

Table 4 Zn and Cd levels in livestock manures and Cd:Zn ratios of the mean Cd and Zn concentrations

Zn Cd Cd:Zn Cu Mean Range Mean Range Mean Mean Range L L L Species mg kg 1 dry matter g g 1 mg kg 1 dry matter Dairy Cattle FYM 153 99e238 0.38 <0.10e0.53 0.0025 37.5 26e56 Dairy Cattle Slurry 209 <5e727 0.33 <0.10e1.74 0.0016 62.3 <1e352 Beef Cattle FYM 81 41e274 0.13 <0.10e0.24 0.0016 16.4 10e28 Beef Cattle Slurry 133 68e235 0.26 0.11e0.53 0.0020 33.2 18e49 Pig FYM 431 206e716 0.37 0.19e0.53 0.00086 374 160e780 Pig Slurry 575 <5e2500 0.30 <0.10e0.84 0.00052 351 <1e807 Broiler/Turkey litter 378 208e473 0.42 0.20e1.16 0.0011 96.8 46e173 Layer Manure 459 350e632 1.06 0.44e2.04 0.0023 64.8 49e75

(Nicholson et al. 1999). uu .Chaney L. Rufus Food Safety Issues for Mineral and Organic Fertilizers 75 and do not purchase P-fertilizers or use biosolids because they don't need P and seek to avoid excessive P accumulation in their soils. Keller and Schulin (2003) took into account different amendment applications on qualitatively different farms to examine the Zn and Cd balance. As shown in Table 5, livestock farms had the highest rate of Cd accumulation of all farm classes. Further, the limited amounts of total biosolids in a jurisdiction means that few farms will be able to apply biosolids as a P and N fertilizer, strongly limiting the Cd accumulation from modern biosolids (now much lower in Cd than in the 1970s). Although biosolids with as high as 3700 mg Cd kg1 DW were reported in the older literature (Sommers, 1977), industrial pretreatment of wastewaters, and reduction in Cd use in products has greatly reduced Cd in biosolids. For example, in a study of Pennsylvania biosolids metals levels over a 22-year period, Stehouwer et al. (2000) showed the decline from median Cd of 6 mg kg1 to 1.5 mg kg2 such that the 90th percentile of biosolids Cd was only 4 mg kg1. And in the 2009 Targeted National Sewage Sludge Survey (US-EPA, 2009), the geometric mean Cd was 1.97 mg kg1 (full range, 0.21 to 11.8 mg kg1), while the mean Cd:Zn was 0.0028 [range 0.0006 to 0.0106 g Cd (g Zn)1]. Thus, except in subsistence rice and tobacco production (Bell et al., 1988), modern biosolids do not cause Cd risk to humans or the environment because of their low Cd:Zn ratio and high Cd adsorption strength (Kukier et al., 2010).

Table 5 Estimated Cd accumulation in soils of Sundgau Canton from different sources, and for different farm types

Arable Dairy& Animal Whole Farming Mixed Husbandary Region L L gha 1 yr 1 Input Fluxes Manure <0.1 0.7 8.2 0.6 Biosolids 0.4 0.1 <0.1 0.1 Commercial 0.6 0.2 <0.1 0.4 Fert. Deposition 2.1 2.1 2.1 2.1 Output Fluxes Crop 0.6 1.9 0.6 1.4 Removal Leaching 0.4 0.4 0.4 0.4 Net Flux: 2.3 0.8 9.1 1.4

(Keller and Schulin, 2003). 76 Rufus L. Chaney

5.4.5. Effect of chloride on cadmium accumulation in crops One of the most remarkable aspects of crop Cd accumulation and food- chain Cd risk is the role of soil chloride in Cd uptake. This effect was first demonstrated by Bingham et al. (1984) in pot studies. Then McLaughlin et al. (1994) found that Cd in potato tubers was more strongly affected by soil chloride levels than other soil factors they evaluated. This was soon confirmed for Cd in sunflower kernels (Li et al., 1995), bread wheat (Weggler-Beaton et al., 2000), and durum wheat (Norvell et al., 2000). McLaughlin and cooperators conducted many studies to better understand this interaction between Cd and chloride. For example, increasing chloride in an unbuffered nutrient solution causes some of the Cd2þ to be converted to CdClþ, lowering the activity of Cd2þ, but increasing uptake of Cd by Swiss chard (Beta vulgaris L. var. cicla)(Smolders and McLaughlin, 1996a). Chard was used in these tests because it is highly tolerant of NaCl compared to many other crops and strongly accumulates phytoavailable Cd. Finally, using a chelating resin-buffered nutrient solution approach which maintained constant activity of free Cd2þ in solution, both dissolved total Cd and Cd uptake by Swiss chard was significantly increased when chloride was added to the solution (Smolders and McLaughlin, 1996b). The conclusion of these studies was that when chloride was present at high levels, Cd2þ activity did not control Cd uptake as previously hypothesized. And with the resin- þ 0 buffered system, it is clear that the CdCl or CdCl2 species must be taken up by roots. Another alternative was that the formation of chloride complexes increases the pool of diffusible Cd in the soil up to the root membrane, and because the chloride complexes can rapidly exchange free Cd2þ, they may maintain higher levels of free Cd2þ at the root cell membrane (Degryse et al., 2006a, 2006b). The pronounced effect of field chloride levels were also identified in fields of durum wheat where soil chloride varied across the landscape (Nor- vell et al., 2000; Wu et al., 2002). Interestingly, increased soil solution sulfate also increases dissolved Cd as the sulfate-complex, but that has little effect on Cd uptake by plants (Bingham et al., 1986; McLaughlin et al., 1998). The formation constant for the sulfate-complex is similar to that of chloride. In the Li et al. (1995) study with sunflower kernels, the same highly calcareous soil series occurred at different locations with different chloride levels but both had high soil solution sulfate. But kernel Cd was more than doubled in fields with higher soil chloride with no effect on yield. And among fields across Australia, Cd in potato was more affected by chloride than by soil total Cd or soil pH, illustrating the overwhelming potential effect of soil chloride. In predicting potato Cd, both chloride in the soil at planting and the irrigation water supply were Food Safety Issues for Mineral and Organic Fertilizers 77 important (McLaughlin et al., 1999). Chloride accumulates in poorly- drained soils from KCl fertilizers and other inputs, and can be considerably increased by geogenic sources (Wu et al., 2002). Particularly in alkaline soils, chloride increases Cd uptake more than other soil factors studied, while in acidic soils chloride is less important apparently because Cd solubility is so much greater that chloride complexation is less significant (Hattori et al., 2006). Taken together, these findings indicate that the CdClþ species must be absorbed by roots, but the specific transporter remains unknown; it is considered unlikely that the Zn transporter plays a role in uptake of Cd from CdClþ such that the overall rate of Cd uptake is increased. Most publications on the effect of chloride on Cd uptake support the view that chloride causes higher dissolved Cd due to formation of chloride complexes (but not more than sulfate), and that the rapid exchange of chloride at the root cell membrane lowers charge and promotes uptake. Clearly diffusion is increased, but the difference between the effects of chloride and sulfate is not fully explained by this model. McLaughlin et al. (1998) suggest 0 2þ that the (CdSO4) complex is taken up equivalent to Cd because increasing complexation of Cd2þ by sulfate did not reduce Cd uptake by chard. Alternatively, Khoshgoftar et al. (2004) reported that increased soil chloride reduced the activity of free Zn2þ at the root, increasing Zn deficiency stress on the plant and promoting Cd uptake by the root Zn transporter. This model for Zn deficiency stress increasing Cd uptake was also found in studies of soil amended with a Cd-contaminated biosolids with low Zn. Liming the soil actually induced Zn deficiency in lettuce and strongly increased Cd uptake by lettuce. Zn fertilization to prevent liming-induced Zn deficiency reduced lettuce Cd from 19.3 to 1.95 mg Cd kg1 DW (Chaney et al., 2006). The increase in crop Cd upon liming is in strong contrast with many studies of the effect of pH on Cd uptake (Chaney, 2010). The best available data show that both Cd and Zn are transported on the root Zn2þ transporter (Hart et al., 2002; 2005). Despite an earlier paper suggesting that the root ferrous transporter (IRT1) transported Cd, Cohen et al. (2004) later clarified that if the activity of Cd2þ and other ions in soil solution are taken into account, the ferrous transporter (IRT1) is not involved in Cd2þ uptake.

5.4.6. Retention of Cd from feedstuffs in tissues of livestock Most crops are consumed by livestock, and little of the ingested Cd remains in the livestock tissues consumed by humans. The EU Cd Risk Assessment (EFSA, 2009) ignores the nearly total exclusion of Cd from the carcass of livestock even when their diets are enriched in Cd. This has been tested in experiments where biosolids were added to cattle diets to simulate the 78 Rufus L. Chaney inadvertent ingestion of biosolids during grazing (Kienholz et al., 1979), or biosolids were used to produce corn silage fed to goats and sheep. In the Kienholz et al. (1979) study, feedlot cattle were fed diets with 0, 4, or 12% biosolids (dry weight basis) for 94 days before slaughter and analysis of tissues. The biosolids contained 21 mg Cd and 1500 mg Zn kg1 DW, higher than modern biosolids after pretreatment has become widely enforced. They measured that <0.04% of diet Cd was retained in the whole carcass of the cattle at the end of the experiment. Table 6 shows the balance of Cd in a three-year experiment with dairy goats reported by Bray et al. (1985). Even with the highest rate of biosolids, and the high Cd concentration and Cd:Zn ratio in the biosolids applied, the goats only retained 1.9 mg Cd in liverþkidney of the 3980 mg of Cd which were in their feed. Hinesly et al. (1985) similarly found a very low fraction of the Cd in corn and soybeans fed to chickens to be retained in the chicken bodies, and none in eggs even for high Cd biosolids. Decker et al. (1980) found no increase in tissue Cd in cattle which grazed on low Cd biosolids amended pastures or were fed low Cd biosolids in diets. Similarly, adding 5 ppm Cd as salt to diets of diary cows did not cause any increase in milk or meat tissues Cd concentration (Smith et al., 1991). Some earlier studies which used addition of Cd salts to livestock diets found accumulation of Cd in tissues, but when Zn is added along with the Cd, much lower accumulation is observed (Stuczynski et al., 2007). In the latter study, forage from uncontaminated soil and from a remediated Zn smelter slag were fed to calves, compared with additions of Cd or CdþZn to the uncontaminated soil grown forage to match levels in the contaminated soil grown forage. The increased Cd in the forage grown on the remediated site caused a small signiﬁcant increase in kidney Cd (0.53 vs. 0.17 mg kg1 wet weight from control forage), while

Table 6 Accumulation of Cd by goats fed corn silage grown on biosolidsa amended soils for three years; silage composition for year-3

Corn Silage Feed Liver Kidney Muscle L D K Cd Zn Cd Cd Cd Cd Cd Treatment mg/kg DW mg/3 yr mg/kg DW mg Control <0.06 34.5 <40 0.26 3.09 <0.06 0.17 Low 1.39 73.5 1290 1.72 10.9 <0.05 0.77 Medium 2.73 74.6 2180 2.10 24.7 <0.07 1.32 High 5.26 107 3980 2.94 22.4 <0.12 1.41 a Biosolids contained 1700 mg Zn kg-1 and 105e186 mg Cd kg-1, with high Cd:Zn. The high Cd and Cd:Zn ratio allowed higher Cd in crops and goat tissues than commonly found for high quality biosolids (Cd source was a Cd-Ni battery factory). (Bray et al., 1985). Food Safety Issues for Mineral and Organic Fertilizers 79 the Cd salt addition caused a large signiﬁcant increase (2.10 mg kg1) which was reduced considerably by also adding the normal Zn which accompanies Cd in such forages grown on contaminated sites (0.78 mg kg1). Thus the normal livestock production feed-chain comprises very strong limitation of Cd transfer from soils and soil amendments to humans.

5.4.7. Exceedance of kidney Cd limits in sheep due to historic P-fertilizer Cd Because the P-fertilizers used in Australia and New Zealand for decades were rich in Cd, soils, forages, and sheep kidney and liver of older animals showed signiﬁcantly higher Cd accumulation than where low Cd fertilizers were used (Roberts et al., 1994; Loganathan et al., 1999). Thus, Australia and New Zealand have set low limits on Cd in kidney and liver of sheep in order to meet import limits in the EU. Factors which cause kidney of older sheep to exceed these limits may clarify if similar outcomes would be expected in other countries. A number of factors interact to create a remarkable Cd transfer case in some Australian and New Zealand sheep. First, the long application of high Cd phosphate fertilizers from Nauru rock phosphate, soil acidity or chloride, and age of the animals are each important. Only kidneys of older animals fail to comply with EU Cd limits. After detailed investigation, it was also discovered that unusual accumulation of Cd but not Zn by capeweed (Arctotheca calendula) [an introduced weedy legume which infests pastures and grows well in winter] causes the total Cd and the Cd:Zn (already high from the high Cd:Zn ratio of phosphate fertilizer) of the sheep diet to be substantially increased compared to animals grazing ﬁelds where capeweed was controlled (McLaughlin et al., 1996, 1997; Hamon et al., 1997; Bramley and Barrow, 1994). Also, liver and kidney limits have been based on assumptions of regular lifetime consumption of liver and kidney rich in Cd. This scenario is unrealistic and generates unnecessarily low Cd limits for kidney and liver, and in fertilizers and other soil amendments.

5.4.8. Has there been an important increase in crop Cd over time? Concern about increasing food Cd over time was initiated by a report of Kjellström et al. (1975) who analyzed wheat samples from different harvest years from a location in Sweden. And increased by the assumption that Cd in applied fertilizers necessarily causes increased Cd in crops. The samples they analyzed indicated that crop Cd had doubled in 50 years, so they expressed concern about this pattern. Andersson and Bingefors (1985) tested more samples from this farm operated by a plant breeding company and found large annual variation in grain Cd over time. The wheat was grown on different fields in different years due to normal crop rotation; different fields had different pH levels and Cd 80 Rufus L. Chaney levels from parent materials, which caused large year-to-year variation in crop Cd concentration. Thus when field-to-field variation and cultivar differences were considered, it became unclear whether a meaningful increase in wheat grain Cd had occurred. There was no increase in barley grain Cd. A more recent comparison of Cd in wheat grown on long-term soil fertility field test plots from 1967 through 2003 showed a significant decline in grain Cd, attributed to reduced aerosol deposition of Cd in Sweden (Kirchmann et al., 2009). Lorenz et al. (1986) conducted a similar evaluation of changes in crop Cd from long-term samples maintained at a location in Germany, but found no increase in wheat Cd over time. Recently Jorhem et al. (2001) tested flour from the period 1983e1997 and also found no significant increase in wheat flour Cd over time. Rothamsted long-term field plots tested the change over time of soil and crop Cd, finding increased grain Cd on very acidic soils (but not if the soils were limed) but decreased grain Cd with considerably higher amounts of manure-applied Cd ( Jones and Johnston, 1989). Further, a long-term comparison of Cd in foods in Finland also found significant decline over time rather than increase (Ekholm et al., 2007). Examination of the balance of Cd, Zn, P, and some other elements in cropping systems has been undertaken by scientists in several countries. By examining the composition of all sources of Cd for soils (fertilizer, irrigation water, aerosol deposition, manure, biosolids, etc.), and estimating the normal rate of nutrient amendment application limited to avoid excessive application of N or P to soils, they could evaluate the role of farming system and soil amendments on long-term Cd balance. The results are quite interesting because they are counterintuitive. With the composition of P-fertilizers being used in 2000, both aerosol deposition and manure applications applied more Cd to farm types of Switzerland (Table 5) than did fertilizers (Keller and Schulin, 2003). Livestock production farms import feed supplements and apply the livestock manure onto their fields, but need no additional P-fertilizer and may need no N-fertilizer. And would apply no biosolids because that would add additional unnecessary P. A similar conclusion was reached in the assessment of long-term accumulation of Cd in soils and crops of Canada (Sheppard et al., 2009).

5.4.9. Geogenic Cd enrichment of cropland Cadmium, arsenic and other potentially problem elements are part of the natural world. Although soil Cd is accompanied by 200e500-fold higher Zn in most soils, a few soils developed from mineralized rocks such as marine shale (Lund et al., 1981) or alum-shale (Mellum et al., 1998) (also called iron sulfate soils) which are Cd mineralized without usual Zn co-mineralization. Soil development in Somerset, UK, included some Zn ore materials which caused higher Cd in wheat grain (Chaudri et al., 1995). Natural Food Safety Issues for Mineral and Organic Fertilizers 81 mineralization of marine shale parent rocks on the western side of Salinas Valley, CA, has provided much Cd but little Zn to residual soils, and Cd- enriched regional alluvial soils (see Holmgren et al., 1993; Chen et al., 2009). Combined with high chloride in irrigation waters, these soils promoted Cd accumulation by leafy vegetables (Burau, 1983; Wolnik et al., 1983; Chen et al., 2009). Alternative land use kept most of the high Cd residual marine shale soil from producing vegetables for decades, but changing economics changed farm practices. Interestingly, liming these soils did not reduce Cd accumulation by spinach (Burau, 1983), in strong contrast with the usual strong reduction in Cd uptake by crops when soil pH is raised by 1 or more units (Chaney, 2010). Chaney et al. (2009) tested the hypothesis that the low level of soil Zn, coupled with high Cd:Zn ratio, caused plants to experience functional Zn deﬁciency on such soils, causing plants to attempt to obtain more soil Zn but absorbing the Cd present as found with a soil very high in Cd from a contaminated biosolids (Chaney et al., 2006). When adequate Zn was applied along with the limestone to make soils with 9 mg Cd kg1 calcareous, lettuce Cd fell from over 12 mg kg1 DW to well under the 4 mg Cd kg1 DW of the proposed Codex standard for leafy vegetables (Chaney et al., 2009). Perhaps the most remarkable example of geogenic Cd enrichment of soils was identiﬁed in Jamaica (Lalor et al., 1998). In some locations, natural soil Cd exceeds soil Zn! Geological research found that marine fossils provided the Cd to the parent rocks millions of years ago that formed these soils (Garrett et al., 2008). Because most of the soils are near neutral pH, very rich in Fe and Mn oxides, and not high in chloride, local crops do not contain extreme levels of Cd, although clearly above background levels (Howe et al., 2005). In contrast with the cases in rice Cd in Asia, clear evidence of renal tubular dysfunction has not been observed in Jamaica (Lalor et al., 2004; Lalor, 2008; Wright et al., 2010). With this example, other areas with similar soil genesis from marine shale sources need to be examined for potential food-chain Cd risk from mineralized soils.

5.4.10. Deliberate contamination of a fertilizer product Although most fertilizer manufacturers conduct careful quality control, a few have violated business norms. One egregious case occurred with a ZnSO4 product marketed by one or more Chinese chemical companies. The situation was discovered when importation of the ZnSO4 was sought in British Columbia, Washington, Oregon and California in 1999. Because of the Seattle newspaper articles (Wilson, 2001), each of these jurisdictions had fertilizer laws and enforcement agencies, so samples of the product had to be locally analyzed and the product registered on that basis. But the product had remarkable levels of Cd (Table 7), far above natural abundance (Chaney, 2010). One sample had as much Cd as Zn, 82 Rufus L. Chaney

Table 7 Cd and Zn concentration and Cd:Zn ratio of in samples of Cd-contaminated Zn fertilizer product delivered to northwestern US and Canada

Cd Zn Cd:Zn L L Sample mg kg 1 DW g g 1 China-Zn-1 46,400 345,000 0.135 China-Zn-2 72,800 313,000 0.233 China-Zn-4 215,000 216,000 0.995 China-Zn-5 199,000 230,000 0.865 a Cenes ZnSO4 7.1 320,000 0.000022 Blue-Minb 49 420,000 0.000127 a Zn fertilizer grade product. b Byproduct Zn fertilizer. (Washington State Department of Agriculture, unpublished data, 1999). indicating that waste Cd was collected from other sources to mix with and adulterate the ZnSO4 ingredient. Subsequently, highly Cd-contaminated ZnSO4 products were sold in other countries. News reports of Cd contamination in South Africa, Kenya, Australia and the EU (France, Belgium) showed that the contaminated ZnSO4 exports continued in the 2000s (ec.europa.eu/food/ committees/regulatory/scfcah/animalnutrition/summary38_en.pdf). In these cases, the vendor certified the product contained low levels of Cd (e.g., <15 mg kg1), but eventual analysis showed much higher Cd levels. Zn product used in animal feeds harmed livestock from the Cd present. And high Cd in the Zn fertilizer or livestock manure applied to crops caused crops to exceed import limits in the EU. Few details of these cases have been published, but the harm is persistent. A single application of this highly contaminated Zn fertilizer may include 2e5kgCdha1 when applying 5e10 kg Zn ha1. This case is instructive. Simply having a good law on trace elements in fertilizers is not enough. These products harmed farmers in Australia and the EU which had clear fertilizer Cd limits. Enforcement of the law such as occurs in Washington State, California and Canada is required for adequate protection to be obtained (see Washington State Department of Agriculture web site). Analysis of commercial fertilizer product samples is required before registration and approval for sale; companies must confirm that their product remains below state limits in trace elements in reports to the WA Department of Agriculture. It is likely that highly Cd-contaminated Zn fertilizer products have been sold in many other countries and parts of China; but without monitoring, no data are available to estimate the extent of any problem. The People's Republic of China (PRC) is reported to have responded to complaints from the EU by agreeing to Food Safety Issues for Mineral and Organic Fertilizers 83 conduct government testing of any Zn fertilizer or feed additive products to be exported to the EU, and a PRC government analysis certificate would accompany any product sold in the EU. That does not protect other nations with fewer regulatory protections. A deliberately Cd-contaminated Zn fertilizer product is especially worrisome in Asia where use of such Zn fertilizer on rice soils (which often receive Zn fertilizer) could eventually cause human Cd disease to be discovered decades later in subsistence rice farm families.

5.5. Lead Because Pb is so strongly adsorbed or precipitated in soil and plant roots, plants do not accumulate high Pb concentrations even when soils are appreciably contaminated with Pb. Pb phytotoxicity does not occur in the general environment (Koeppe, 1981) even when soil Pb levels comprise risk to children. Urban soils are remarkably contaminated by historic automotive emissions, stack emissions, and house paint residues, particularly in the inner city (Mielke et al., 1983). Because Pb uptake into edible parts of plants is so small, soil ingestion is the most limiting risk from soil Pb, and high Pb soils are not recommended for gardens to limit soil ingestion exposure. But urban citizens commonly grow garden crops in such high Pb gardens around the world. Fortunately, application of phosphate fertilizers and compost products can strongly reduce the bioavailability of soil Pb to persons who ingest soil (Sterrett et al., 1996; Zia et al., 2011). A field test was conducted at Joplin, MO, near a Pb smelter where contaminated lawn soil containing about 3000 mg Pb and Zn kg1 was treated with different soil amendments to attempt to reduce soil Pb risks. The study included feeding untreated and P-treated soils to human volunteers. Soil Pb bioavailability was reduced 69% by treatment with phosphoric acid, nearly as much by treatment with triplesuperphosphate, but less with rock phosphate (Ryan et al., 2004). Soil Pb was converted to chloro-pyromorphite, a form of Pb with very low bioavailability (Scheckel and Ryan, 2004). When high applications of phosphate are to be used to reduce soil Pb risk, products low in Cd should be selected to avoid the possible high Cd loading with a remediation application of phosphate. Lead arsenate pesticide was major source of Pb contamination of orchard soils (Peryea and Creger, 1994; Merwin et al., 1994). Numerous studies of Pb uptake by garden crops found little uptake (e.g., Elfving et al., 1978). But the US Food and Drug Administration found two samples of commercial carrots with anomalous levels of Pb, and advised the industry to find a way to prevent this from occurring again. The industry contracts for most carrot production, so they required growers to prove that their soils were not high in Pb, much like carrot growers 84 Rufus L. Chaney have long been required to show their soils are not high in DDT. Codling et al. (2012) tested why carrots contained high Pb levels when grown in old orchard soils and found that the Pb was not soil particles adhering to the carrot peel layer, or Pb ions accumulated into the peel layer cells, but was within the peeled carrot. Chaney et al. (2010a,b,c) examined the localization of Pb in these carrots using m-X-ray fluorescence and found that the Pb was within the xylem. The xylem of root crops (which are actually expanded hypocotyls) grows through the edible root, so part of the absorbed Pb moving to the shoots is trapped in the xylem. In contrast, potato tubers are phloem fed and contain no increase in Pb when grown on the same Pb-rich orchard soils (Codling et al., 2012).

6. Trace Elements in Other Common Soil Amendments

6.1. Manure All livestock manure contains Cd, Cu, Zn, etc. from the feed ingredients and P-supplements. Nicholson et al. (1999) summarized the Cd, Zn, Cu, As and other element concentrations in manure from different livestock classes in the UK. The means and range for Zn, Cd, Cu, and the Cd:Zn ratio of the means are shown in Table 3. These data show that swine and poultry feeds are commonly supplemented with high levels of Zn and Cu, and may contain an As compound as a growth stimulant or antibiotic. Because of the high levels of Zn and Cu in swine manure, repeated manure applications are the dominant source of increase in soil Zn and Cu in many nations (Sheppard et al., 2009; Keller and Schulin, 2003). Accumulation of As in soils amended with poultry litter has been observed in some experiments (Adeli et al., 2007), while not in others (Codling et al., 2008). Because manure contains high phosphate levels, and phosphate competes more strongly than arsenate for adsorption sites in soils, the co-applied phosphate can cause leaching of applied As. So for the low Fe, light-textured Coastal Plain soils in the Codling et al. (2008) experiment, As was not increased after decades of poultry litter application. But in the Adeli et al. (2007) study, and other research with heavy textured and/or high Fe soils, poultry litter applied As remains in the plow layer depth (as do As from orchard sprays, herbicides and defoliants) (Staed et al., 2009). Despite low levels of Cu in normal forages used as ruminant feedstuffs, McBride and Spiers (2001) found high Cu in dairy manure. The source was found to be CuSO4 use in hoof baths to prevent a hoof disease in dairy cattle. The CuSO4 bath becomes fouled with manure and soil over Food Safety Issues for Mineral and Organic Fertilizers 85 time, and was commonly dumped into the liquid manure handling system. Dairy manure with 1000 to even 3000 mg Cu kg1 DW have been reported. Tests have shown that livestock producers do not need to use the Cu baths continuously; and the fouled baths can be disposed with other solid wastes rather than dumped into the manure. Producers are presently advised to minimize manure Cu to prevent future problems with Cu accumulation in soils. Adverse effects of manure Zn and Cu have seldom been reported perhaps because the manure adds metal sorbent to amended soils, and normal near neutral pH management minimizes the potential for Zn or Cu phytotoxicity. All phytotoxicity of soil Zn, Cu and Ni occur at strongly acidic pH until extreme amounts of metals are applied, far more than in present manures (Chaney, 1993, 2010). Historically excessive application of Cu pesticides and fertilizers in Florida, USA, caused accumulation of Cu in sandy soils low in organic matter, and with low pH. These conditions promoted Cu phytotoxicity to citrus and vegetables (Reuther and Smith, 1954), which was counteracted by liming and adding FeEDTA to correct Fe deficiency induced by the Cu. In European long-term vineyards, applications of Cu pesticides caused accumulation of Cu to high levels such that some grasses suffered Cu toxicity (Cu-induced Fe deficiency) even at neutral pH which should ordinarily prevent Cu phytotoxicity. In this case, the phytosiderophores secreted by young roots of durum wheat were filled by Cu rather than Fe3þ, thus inhibiting the normal Fe uptake mechanism of grasses (Michaud et al., 2007). Similar effects on dicots are not observed in these soils, but the chlorosis of durum wheat is an adverse effect of historic Cu pesticides in vineyards and some excessively Cu fertilized soils. Concern about As in soils and land-applied poultry litter are discussed below. The tolerance of increased soil As is very complex, and the benefit of arsenicals in poultry diets have been demonstrated ( Jones, 2007). False claims about As retention in chicken carcasses (Lasky et al., 2004) have confused many about transfer of As to humans in the poultry meat consumed (Tsuji et al., 2007). But objective measurements have shown that poultry fed arsenicals have little or no increase in tissue As except in liver, and that is depleted during the required withdrawal period. So the most direct justification to stop use of arsenicals may lack a technical basis. As is described below, extreme claims about cancer risk from exposure to even below-background levels of As in soils, water and food suggest that even background levels of soil As cannot be tolerated. It seems unlikely for this to occur in nature.

6.2. Limestone Byproducts containing limestone equivalent have often been recovered during Zn and Pb ore beneficiation, and historically these materials were 86 Rufus L. Chaney used as agricultural limestone. Some other byproduct limestones are quite low in trace elements (sugarbeet lime; acetylene lime). Chaney et al. (1988) reported that dolomitic limestone sold in the vicinity of Palmerton, PA, had usually been the “Horsehead” brand, a product of Zn ore beneficiation at the Friedensville, PA, mine of the New Jersey Zinc Company. Three samples contained 3.6, 3.9 and 8.4 mg Cd kg1 dry weight, and 2500, 2810 and 4580 mg Zn kg1 (Cd:Zn ratios are 0.00156, 0.00139 and 0.00173, all far below the ratios of ores and fertilizers). Normal agricultural limestone contains <0.3 mg Cd kg1 and 21 mg Zn kg1. Other Information indicated that a similar Cd and Zn- contaminated mining by-product limestone from Tennessee was sold in several Southeastern states. Pb, Zn and Cd rich limestone byproducts were often marketed in Mis- souri, USA, and other states from the mines in the “tri-state area mining district” (KS, OK and MO). Several experiments were conducted to test the effect of using these products on trace element accumulation by garden crops (Davies et al., 1993; Wixson et al., 1984). As might be expected, use of limestone raised soil pH and reduced the phytoavailability of soil Zn and Cd very strongly. But application of a limestone product with 10 mg Cd kg1 at five t ha1 every five years would add 1.0 kg Cd ha1 every century, more than estimated for application of P-fertilizers.

6.3. Steel production fume waste In a case in Tifton, Georgia, USA, a steel mill fume waste rich in Zn and limestone equivalent was sold as a substitute limestone source. The combination of light-textured soils and production of peanuts led to severe Zn phytotoxicity in peanuts grown on these soils several years later as the soil pH naturally fell due to rainfall and fertilizer use (Davis et al., 1995). Peanut is unusually sensitive to excessive Zn in acidic soils (Keisling et al., 1977). Fortunately, high Zn in the soil amendment restricts Cd accumulation in a peanut (data not reported for the Georgia case) as shown by McLaughlin et al. (2000).

6.4. Gypsum

Gypsum (CaSO4$2H2O)is applied to soils as a Ca or sulfate fertilizer, as an soil aggregation aid to water infiltration, to prevent crusting, and to reduce erosion; to aid in alleviation of subsoil Al phytotoxicity, and as a Ca source to remediate sodic soils (contain excessive Na). In the past, mined gypsum and phosphogypsum (from the phosphate fertilizer manufacturing industry) were used in many nations. Mined gypsum does not commonly contain trace elements at levels which comprise concern. Phosphogypsum may contain Cd and Zn, and Ra and other isotopes which could limit use of Food Safety Issues for Mineral and Organic Fertilizers 87 this form of gypsum on land (Mays and Mortvedt, 1986; Rutherford et al., 1994). Phosphogypsum contains a variable fraction of Cd from the phosphate rock depending on the manufacturing process (see Wakefield, 1980). More recently gypsum manufactured during removal of SO2 from the exhaust of coal-fired power plants (Fluidized Gas Desulfurization Gypsum, FGDG) has been evaluated for beneficial use in agriculture. Until the 2000s, this gas treatment was conducted with fly ash removal and some of those products contained potentially phytotoxic levels of boron, and could have high levels of Se, As, etc. (e.g., Kost et al., 2005; Ransome and Dowdy, 1987). But when it became evident that the trace element contaminated FGDG would not be acceptable in agriculture, the industry tested changing the gas treatment train to remove fly ash first, then to remove the SO2 and produce clean FGDG. Power plants are required to remove the SO2, and by 2020 are expected to generate on the order of 30 million t yr1 in the US. Fortunately this change to much cleaner FGDG has prevented unusual levels of nearly all trace elements from entering the FGDG. Only Hg, Se and As were somewhat higher than background soils, while Cd, Zn, and other trace elements are similar to soils and mined gypsum (Chaney et al., 2010). In the presence of the huge 2- 2- amount of SO4 , SeO4 uptake by crops is not increased, so the slight increase in Se in the FGDG does not comprise risk. Sulfate poisoning of cattle was observed when appreciable levels of gypsum were added to a feed supplement for range cattle (Raisbeck, 1982). Excessive sulfate intake by ruminants can cause a severe disease, polioencephalomalacia (Gould, 1998). Adherence of powdered gypsum on leaves following surface application on pastures could allow increased intake by grazing livestock, as could direct access to FGDG piles in a field. Thus, a waiting period for rainfall or forage growth should occur after surface application of powdered gypsum before ruminants may graze amended pastures, and ruminants should be fenced out of FGDG storage piles.

7. Long-Term Reactions of Trace Elements in Soils

It has long been known that the Zn in ZnSO4 fertilizers added to deficient soils was slowly transformed to forms with lower phytoavailability (e.g., Boawn, 1976). A normal fertilizer application on a deficient soil may require additional Zn fertilization after 5 or more years. The mechanism of Zn becoming less phytoavailable over time has recently become better understood. Initially, it was thought that the added Zn became more strongly bound on Fe and Mn oxides, or diffused into micropores in the 88 Rufus L. Chaney oxides where it could not reach soil solution easily (Bruemmer et al., 1988). Since then research has found that new Zn minerals are formed in soil, the ZneAl-layered double hydroxide (LDH), which may then gain silicate and become even less phytoavailable (Scheidegger et al., 1997). Formation of the Zn-LDH is favored at higher soil pH. Zn in long-term Zn- contaminated soils is largely transformed to the Zn-LDH (Voegelin et al., 2005). Ni, Zn and Co form LDH species in contaminated soils (Voegelin et al., 2002), but Cd does not, helping to explain why Cd can remain adsorbed, labile and phytoavailable for decades after addition to soils (e.g., Kukier et al., 2010). There is some evidence that fertilizer- applied Cd becomes somewhat less phytoavailable over time (Hamon et al., 1998), but to a far lesser extent than Zn. One of the most important errors in research to better understand long- term risks from soil trace elements was considering that freshly added soluble metal salts represented long-term risk. Besides the long-term reactions of adsorbed metals to become less phytoavailable (hysterisis), other errors were identified. When a soluble salt, say ZnSO4, is added to soil, the Zn reacts with the soil surfaces and displaces protons causing lower pH and higher Zn phytoavailability (White et al., 1979; Speir et al., 1999). Large additions of Zn caused reduction in soil pH of several units which greatly increased the phytoavailability/phytotoxicity of the added Zn. Even if the soil is kept at the same pH by additions of CaCO3, until the anion is leached from the soil the presence of the accompanying anion causes soil solution to contain higher levels of Ca, Mg, etc. These compete with Zn for sorption sites and keep the activity of free Zn2þ in the soil solution higher. Thus until amended soils are leached to remove the accompanying anion, phytoavailability is higher whatever the pH. And until time has transpired to allow the Zn to move into new solid phases and sorption equilibria to be reached, Zn is much more phytoavailable than it becomes over time. Some of the best evidence to illustrate this point comes from study of Zn salt additions to soils collected near and far from galvanized steel towers for electricity lines. Near the towers, soils can reach 10,000 mg Zn kg1.In strongly acidic soils this high Zn inhibits nitrification and causes phytotoxicity (which caused the selection/evolution of Zn tolerant grasses) (Al-Hiyaly et al., 1988). But when soil pH is near 6.5 or above for the very high Zn soils, no adverse effects are seen, not on plants or on soil microbes (Smolders et al. 2004, 2009). If equivalent Zn salt is added to the “control” soil collected far from the tower, and pH kept equivalent, severe toxicity to soil microbes was observed. Thus the extreme concern expressed by some about accumulation of Zn in soils being a threat to soil microbes is now seen as an artifact of the use of soluble Zn salts in study of microbial toxicity. These conclusions apply to all studies in which soluble metal salts are added to estimate effects of metals in soils. Food Safety Issues for Mineral and Organic Fertilizers 89

8. Other Elements in Fertilizers and Soil Amendments of Possible Concern

Other elements occur in particular soil amendments which have raised concern about potential phytotoxicity or food-chain risk. These are considered below. In phosphates, F, As and U have raised concerns and risk assessment research has clariﬁed the basis for risk.

8.1. Fluoride Phosphate fertilizers commonly contain considerable amounts of F (1e3% in superphosphate), and most of the F must be removed from P sources before such materials can be used as defluorinated phosphate feed additives. Excessive F absorption from contaminated feedstuffs or feed supplements by livestock interferes with bone and teeth health (fluorosis) (Suttie, 1980). The NRC (1980) suggests that the maximum allowable level of F in ruminant diets is about 40 mg kg1 for the most sensitive grazing animal classes (bovines). Fortunately, F is strongly bound or precipitated in soils, is sufficiently poorly translocated to plant shoots, and is so phytotoxic that shoots of plants injured by soil F absorption do not exceed 40 mg kg1 (Stevens et al., 2000). Most F poisoning of livestock occurred where crops were exposed to and accumulated high levels of F from air pollution of aluminum smelters, etc. (Weinstein, 1977; Weinstein and Davison, 2003). As might be expected, the bioavailability of different forms of F in diets vary considerably, with NaF more bioavailable than CaF2 than phosphate supplements (Shupe et al., 1962; Ammerman et al., 1964). Aluminum smelters or other industrial emissions commonly caused plant foliage to contain dangerous levels of F, but when orchardgrass was grown on the contaminated soil in a clean air greenhouse, the plants contained only basal levels of F (Braen and Weinstein, 1985). Besides aluminum smelters, phosphate fertilizer manufacturers, brick works, glass manufacturers, and electronics manufacturers commonly emit HF or SiF4 gases to the local atmosphere which can poison plants which can impact livestock and wildlife. Some steel smelting emits F, and downwind vegetation was highly F-enriched at a site in Utah. The forage F was fed to cattle in comparison with NaF and CaF2, and the forage F was found to be as toxic and bioavailable as NaF (Shupe et al., 1962). Concern about F uptake by forages grown on biosolids amended soils led to testing by Davis (1980) when a UK biosolids with 33500 mg F kg1 was identified. Most biosolids contained only as high F as fertilized soils (about 300 mg kg1)(Rea, 1979). Perennial ryegrass was grown on control and biosolids amended soil immediately after the biosolids were 90 Rufus L. Chaney incorporated into the soil. Fluoride accumulation by three clippings of perennial ryegrass declined rapidly to near background levels after incorporation of the biosolids into the test soil. However, the potential for direct ingestion of such high F in biosolids amended surface soil, or biosolids adhering on plants, raises concerns. Chaney and Lloyd (1979) showed that fluid biosolids spread out on, dried on, and adhered to leaves unless immediately washed off. Adhering F-rich biosolids or manures could carry excessive F into diets of grazing livestock even when plants cannot accumulate excessive F from amended soils. Based on the evidence of livestock ingestion of biosolids adhering to forages, Kienholz et al. (1979) fed biosolids at 4 and 12% of diet to cattle to examine the bioavailability of biosolids trace elements and xenobiotics to livestock. Although the increase in bone F was not extreme and no adverse effects of the F were observed, animals fed biosolids did have significantly higher bone F than the control animals, and in that case, the F concentration in the biosolids was more typical at about 300 mg kg1. Thus F-rich biosolids (enriched due to discharge by specific industries) should be injected or incorporated into soils and crop established before grazing livestock are exposed to a field. Highly F enriched biosolids should not be used in agriculture. To date, F in biosolids has not been regulated in the U.S. despite this evidence being provided to US-EPA. Risk assessment for F in phosphate fertilized soils has been conducted by several groups (Cronin et al., 2000; McLaughlin et al., 2001; Stewart et al., 1974; Loganathan et al., 2001, 2003, 2008; Grace et al., 2003, 2005). The initial risk assessments concluded that fertilizer-applied F could theoretically cause adverse effects in grazing cattle and sheep, and appreciable field research was conducted to support improved risk assessment. Eventually, soil feeding tests were conducted to assess soil-F bioavailability. Fluoride applied in phosphate fertilizers quickly reacts with soils to reach low phytoavailability (Loganathan et al., 2001; McLaughlin et al., 2001). Forages grown in fields with high cumulative loadings of phosphate borne F show a small increase in shoot F, but do not reach phytotoxic or zootoxic F levels in forages. One important aspect of the potential risk from fluoride ingestion by grazing livestock is the adherence of particles of the applied phosphate fertilizer on the surfaces of leaves (Stewart et al., 1974). Pasture herbage collected the day fertilizer was broadcast on the pasture contained up to 800 mg F kg1, while unfertilized herbage contained < 2mgFkg1. By about 30 days post fertilization, F levels in the herbage dropped to background levels. Thus potential direct ingestion of the P-fertilizer adhering to leaves may require a livestock withholding period for P-fertilized pastures. In some field tests, when livestock grazed pastures with varied rates of phosphate fertilizer, the Food Safety Issues for Mineral and Organic Fertilizers 91 bone F of sheep increased significantly (but mildly) with increasing phosphate rate (Mason et al., 1989; Stewart et al., 1974). Now these experiments can be interpreted as direct ingestion of the phosphate fertilizer by the grazing animals which were not removed from the pastures during the experiment. But ingestion of soil by grazing livestock is well known, and the bioavailability of F in ingested surface soil of P-fertilized pastures (rich in F) was initially assumed to be high. Additional research evaluated the bioavailability of F in soils ingested by livestock, both in the field and in controlled feeding studies. Results have varied among studies from quite low to moderately high, based on different approaches to assessing “bioavailability” (e.g., Milhaud et al., 1989 vs. Grace et al., 2003). In the end, accumulation of F in bone is the key measure of risk from ingested F, not the net absorption of F from diets as reflected in urine F or balance studies. As noted above, CaF2 is substantially less bioavailable than NaF, so it should be no surprise that soil-F with its low solubility due to reaction of F with the soil minerals would have low F-bioavailability. Study of serum F in soil fed livestock indicated that until serum exceeded about 0.25 mg F L1, no increase in bone F resulted. Urine F was increased more than bone F by soil feeding compared to NaF ingestion (Grace et al., 2003, 2005). Thus, soils with common levels of accumulated phosphate fertilizer- F cause little increase in bone F in cattle or sheep. But crops enriched in soluble F by air pollution cause remarkable F accumulation in the animals and F disease. Even with soil containing 1000 mg F kg1, no risk was evident from feeding research. Average soil ingestion is only about 2.5% of diet, while maximum short term soil ingestion reaches 15% of diet or higher (during winter with short pasture or over-stocked pastures). So the timing of exposure is important, especially in New Zealand research where winter pastures with inadequate forage causes increased soil ingestion (Grace et al., 2007). During periods when soil-F ingestion is low, some previously absorbed F is excreted by grazing livestock, so the long-term soil-F ingestion pattern shows lower risk than estimated from short term feeding studies at winter soil intake rates (Grace et al., 2007). Continuous exposure to high F soils is required to show clear evidence of higher F absorption, while livestock are rotated among pastures to maintain adequate forage supply and graze in seasons other than winter. And even with soils with relatively high accumulated F from high end phosphate fertilizer applications, little F was absorbed and deposited in bone during soil feeding tests. Because of the reactions of F in soils, plants and livestock, foods hardly cause any increase in exposure to F from soils with significant F enrichment except by soil ingestion. At many industrial/mining or F-mineralized sites, livestock and wildlife may be at risk from soil F (e.g., Geeson et al., 1998; 92 Rufus L. Chaney

Andrews and Cooke, 1989). Wildlife have suffered adverse effects on teeth and bones when exposed to high F soils even when plants accumulate little F internally (Schroder et al., 1999, 2003). Splash of high F soils onto plants or trampling of forages into soils by grazing livestock increase soil ingestion- exposure of livestock and wildlife. An illustrative case of F poisoning of humans was identiﬁed in China where high F coal and charcoal were used to dry crops or heat homes in open ﬁres which transferred ash particles very rich in F to foods and humans (Finkelman et al., 1999) Also, a few plant species strongly bioaccumulate F from soil; one of these is tea (Camilla sinensis). Several risk assessments have been conducted for F exposure from tea consumption (e.g., Lung et al., 2008). Phosphate fertilizers for tea probably should contain low levels of F to minimize F accumulation in tea.

8.2. Radionuclides Radionuclides (226Ra, 238U, 228Po and related isotopes) are commonly present in most phosphate rock ore (especially from marine sources compared to volcanic sources) (Menzel, 1968; Mortvedt, 1994). The radionuclides are distributed between products and wastes during processing with more 228Ra going to the phosphogypsum than the fertilizer. Electric furnace phosphate can be nearly devoid of radionuclides, while simple superphosphate contains most of the radionuclides in the rock. Phosphate fertilizers may increase soil U levels depending on their source and processing. Several kinds of risks from the radionuclides in fertilizers have been evaluated. Because 40K is a part of all natural K on Earth, humans receive radiation dose from the K in their bodies and dwellings. Risk assessments have been conducted for workers in the K fertilizer industry. The main conclusion of the 40K assessments is that one should not live in a house within a K fertilizer factory. Because full crop yield requires appropriate K fertilizer, and the K is a natural radiation source, all food has some 40K. This is part of the background natural radiation dose of all on Earth. Fertilizers have little or no effect on 40K dose to humans; it is considered unavoidable and part of the normal and acceptable background radiation dose. Essentially all phosphates contain some U because phosphate ore bodies collect U from leaching groundwater (Menzel, 1968). Thus, phosphate fertilizers apply U to soils, and the accumulation of U in long-term soil fertility plots has been demonstrated (Rothbaum et al., 1979). Because of adsorption on the soil solid phases, U is effectively retained in the tillage depth. Because P fertilization increases U sorption by soils, adding phosphate fertilizer often decreases U concentrations in plants. Food Safety Issues for Mineral and Organic Fertilizers 93

Phosphogypsum storage piles, phosphate ore bodies and wastes, and P-fertilizer storage facilities generate radon gas (222Rn) which decomposes to 210Pb and comprises an avoidable radiation source. Most focus on 222Rn risk is movement of geological 222Rn into dwellings, decomposition of the 222Rn into 210Pb, and exposure of humans through inhalation. Some 222Rn dissolved in well water can provide human exposure through off- gassing within homes, and drinking water (Milvy and Cothern, 1989). Surface water supplies contain much lower levels of 222Rn because it is degassed before or during water treatment. In areas with naturally high groundwater Ra, much of the Ra can be removed during water treatment to produce drinking water. Unfortunately, removal of the water Ra into the drinking water treatment residue makes this residue unacceptable for beneficial use in reducing phosphate in runoff of manured fields (Ippolito et al, 2011). It is estimated that 222Rn contributes about half of the background radiation to which humans are naturally exposed, and a local source of indoor air 222Rn (geological or groundwater) can substantially increase that risk (Cothern, 1999). Atmospheric 222Rn also decays to 210Pb which continuously falls onto Earth, including a minor contamination of all plant surfaces. The 210Pb from atmospheric sources is not considered a significant part of the natural background radiation exposures on Earth. But the 222Rn confined to a dwelling allows higher exposures which should be minimized. There is little evidence that fertilizer use contributes to Rn or Ra in groundwater, while phosphogypsum storage pile leachates 226 can contribute Ra, but this is limited by the solubility of RaSO4 (Burnett and Elzerman, 2001), while use of high rates of phosphogypsum on cropland did not contribute to risk to leachate (Alcordo et al., 1999). Regulations on phosphogypsum are based on preventing excessive exposure to radon. U is very weakly absorbed by plants, even those growing in uranium mine wastes. Radium in fertilizer products is somewhat mobile in crops (Mortvedt, 1994). But little Ra moves to phloem fed storage tissues (seeds, fruits, storage roots and tubers) (Mays and Mortvedt, 1986). Leafy crops can accumulate some Ra from uranium or phosphate mine waste contaminated soils (Million et al., 1994; Tracy et al., 1983), but there is no evidence that normal fertilizer use increases Ra in foods (Sheppard et al., 1989). One interesting comparison in the literature is the evaluation of radionuclides in lettuce grown with nutrient solutions, or in manure fertilized and conventionally fertilized crops. Although much more U was added with the conventional fertilizer, fertilized lettuce radionuclides were no higher than the manured crop, and those were only slightly higher than the nutrient solution grown crop (Lauria et al., 2009). Lack of uptake and translocation of the fertilizer-derived radionuclides is evident in these results as well as those cited above. 94 Rufus L. Chaney

8.3. Arsenic A highly As and Pb contaminated Fe fertilizer was noted above (Williams et al., 2006), but the only common fertilizer with As concentrations higher than background soils are some phosphates. As levels in superphosphate range widely from 3e21 mg kg1 (Table 1)(Chen et al., 2007)orup to 300 mg kg1 in other fertilizers, while background US soils vary from <1.0e18 mg kg1 (2e12 mg kg1 for the 5the95th percentiles; Smith et al., 2005). Some livestock manure contains substantial levels of As from feed additives to promote growth or limit disease such that poultry litter and swine manure may commonly contain 25 mg As kg1 DW (Nicholson et al., 1999). High soil As may also occur where Pb-arsenate was applied to orchards as a pesticide, or where arsenical herbicides or defoliants were repeatedly applied (Staed et al., 2009). The chemistry of As in soils is complicated by unequal adsorption of arsenate and arsenite, and the normal reduction of sorbed arsenate to arsenite in flooded soils. Weaker adsorption of arsenite than arsenate causes higher dissolved As in flooded soils, and hence higher potential for As uptake by rice than crops grown in aerobic soils. Rice is the crop species most sensitive to excessive soil As, which inhibits fertilization of flowers causing empty seeds (called straighthead disease) (Rahman et al., 2008). Interestingly, arsenate is normally absorbed by the phosphate transporter in roots (Zhao et al., 2009), but arsenite uptake is by a silicate transporter (Ma et al., 2008). Because As is so weakly accumulated by plants (Zhao et al., 2009, 2010), livestock and humans are principally at risk from As in ingested soil where the soil was contaminated by industrial pollution or geogenic enrichment. However, serious adverse human health effects of As in drinking water occurred in Taiwan, Bangladesh and West Bengal where drinking, cooking and irrigation water was highly enriched in As from deep geological materials rich in As (Kim et al., 2007). If flooded rice is irrigated with water containing high arsenate concentration, shoot and grain As are considerably increased (Abedin et al., 2002; Hossain et al., 2008). But if rice is grown on As enriched soils with low As irrigation water, rice is not so increased in As. And rice grown aerobically rather than by flooded culture accumulates significantly lower levels of As in leaves and grain (Arao et al., 2009; Li et al., 2009). Silicate amendment may also reduce As uptake by competition with arsenite (Li et al., 2009). As accumulation in rice grain has become a significant issue because risk estimations by several Agencies have led to suggestions that even background levels of As in uncontaminated soil comprise risk to children exposed to the soil. Some suggest that inorganic As in normal rice grown on uncontaminated soils is so high that efforts should be made to breed rice Food Safety Issues for Mineral and Organic Fertilizers 95 with lower As accumulation (Meharg and Raab, 2010; Zhao et al., 2010) or to advise against rice consumption. It does not seem likely that the most important human food crop on Earth (rice) which accumulates levels of only 0.2e0.4 mg As kg1 in grain when grown on uncontaminated soils, should cause adverse health effects. Accumulation of As in the leaves and stem of rice is much higher than in the grain, so that rice stover might need to be avoided as livestock feed for long-lived breeding animals, but normal meat animals will not accumulate unusual levels of tissue inorganic As from such feedstuffs so the food-chain is protected from substantial increase in feed As. One interesting example of a problem feedstuffs and soil amendment rich in As is seaweed (Feldmann et al., 2000; Hansen et al., 2003). Where seaweed is available to farmers at low cost, it is often used as a feed or organic fertilizer. But because of the high As levels present in many seaweed species (Laminaria hyperborea and L. digitata averaged 74 mg As kg1 DW) (Hansen et al., 2003), seaweed and manure from seaweed fed livestock should not be applied to soils. Organic forms of As in seaweed are rapidly converted to inorganic As in aerobic soils (Castlehouse et al., 2003), and may be bioavailable if the soil is ingested. Some fish and shellfish contain high levels of As mostly as arsenobetaine. This organo-As chemical is believed to comprise no risk to consumers because it is absorbed and excreted unchanged and does not cause adverse effects in the body (Vahter et al., 1983). But some of the As in these foods is inorganic (Buchet et al., 1994), so ingestion of fish and shellfish may provide more inorganic As than the vegetable foods (except rice) in the normal diet. Interestingly, some cultures ingest significant amounts of seaweed or cooked seaweed (e.g., hijiki which provides 23 mg inorganic As per meal; Nakamura et al., 2008). Because of As accumulation in seaweed, fish and shellfish, composted residues of these foods may need to be avoided for land application to limit adding As to soils. Another possible aspect of As in agriculture is the contamination of soil by As in CCA-treated lumber used to construct raised beds for gardening (Rahman et al., 2004). Although the As, Cu and Cr do not leach far from the wood, soil within the 0e2 cm zone from the wood surface with original 3e10 mg As kg1 was raised to 50 mg As kg1 after a few years. Contaminated soil from the 0e2 cm zone was used to grow test garden crops and As was slightly but significantly increased in the edible portions of carrot, spinach, etc. Carrot peel had higher As than the peeled carrot. Spinach reached 1.48 mg As kg1 DW, while control spinach contained only about 0.070 mg kg1 DW. Although the increase in crop As grown in soil very near the treated wood is appreciable, the principal As exposure from CCA treated wood is soil/ dust ingestion by children who touch the wood or the soil adjacent to the wood. 96 Rufus L. Chaney

Unfortunately, the risk assessments conducted by US-EPA have used faulty data in the dose response estimation for cancer risk (see Chappell et al., 1997). In Taiwan, residents of a village might have consumed low As water from a surface well, or very high As water from a deep artesian well. But when the cancer incidence epidemiological evaluation was conducted, the data for well water As for different persons was not available. So they used the village median well As concentration to classify all persons in a village, miscategorizing persons with cancer in the village who could have had either high or low water As. Other ﬂaws in the epidemiology have been reported (Chappell et al., 1997; Lamm et al., 2006). In the US, several studies have examined persons who consumed water with moderately high As for decades with no evidence of the bladder cancers caused by excessive As from highly contaminated water in Taiwan (Meliker et al., 2010; Lamm et al., 2004). This is strong evidence that a threshold exists for water As concentration below which no cancer risk exists. Intense policy debate about the risk from low level As exposure continues. During 2010, US-EPA proposed to increase the cancer slope factor for As by 17-fold (US-EPA, 2010). This would reduce the allowed As in drinking water to levels which cannot practically be reached. But worse, the cancer based Soil Screening Level which has been 0.43 mg kg1 (already below background soil As levels) would be lowered to 0.025 mg kg1, far below the natural level in all soils (normal 5the95th percentile range, 2e12 mg kg1)(Smith et al., 2005). Such low possible limits are not a problem for US drinking water regulations because they set the “Maximum Contaminant Level Goal” at 0 mgAs L1, and then set a regulatory limit based on practical matters of achieving and measuring the water As. Public concern about historic soil As contamination risk has been raised by increasing knowledge about soil As from gold mine wastes, Cu and Sn mine wastes, pesticides, defoliants, CCA-treated wood products and other As sources. On the other hand, if soils are rich in amorphous Fe, the bioavailability of soil-As is considerably reduced by formation of FeAsO4 2- (scorodite) or AsO4 adsorbed on Fe oxides (Scheckel et al., 2009; Meunier et al., 2010; Beak et al., 2006). The implications for As in fertilizers and soil amendments of such high cancer slope factors for As are complex. Final decisions may not be reached for years because of lawsuits ﬁled to challenge the change in cancer slope factor. All phosphates contain some As, and if the soil As limit is below background, As in any fertilizer product could be challenged. Such a conclusion seems irrational; increasing soil As by a few percent over decades of fertilizer application can hardly comprise the increase in risk being suggested by toxicologists. In two cases in the US, aerosol emissions of As from industry caused local soils where children lived to become strongly contaminated with Food Safety Issues for Mineral and Organic Fertilizers 97

As. In order to evaluate the potential risk to these children, their exposures were assessed, local soil As levels measured, and speciated As in children's urine was analyzed to assess actual individual inorganic As intakes (Hwang et al., 1997; Tsuji et al., 2005). Until soil exceeded 40e100 mg As kg1, there was no evidence of increased inorganic As ingestion based on As species in children's urine. Whether this low As in urine should be attributed to lower soil ingestion by the children than assumed, or lower bioavailability of soil As, remains unknown. Several soil As bioavailability feeding tests have been reported using monkeys (Freeman et al., 1995; Roberts et al., 2007; Roberts et al., 2002) or pigs ( Juhasz et al., 2007; Rodriguez et al., 2003) which showed that soil As bioavailability was typi- cally 15% of total, far lower than the 100% bioavailable assumed in current risk assessments. For some soils rich in amorphous Fe oxides, bioavailability has been as low as 2% of total As.

9. Monitoring and Control of Trace Elements in Mineral Fertilizers

Based on the experiences and research summarized above, it is clear that fertilizer and soil amendment products with substantial trace element (Cd, As, Zn, Cu, Pb, Mo, Se) enrichment may not be acceptable for long-term use on agricultural soils. After the 1997 Seattle Times articles about contaminated fertilizers (see Wilson, 2001), many jurisdictions developed limits for metals in fertilizers. Several quite different approaches have been used, to varied effect in reducing Cd, Pb and As in fertilizers. Many felt that Cd limits should be based on the P (or P2O5) levels in the product because most of the Cd in P and mixed fertilizers comes from rock phosphate used to produce the fertilizer. But research has shown that high Cd products may cause clear increase in crop Cd over decades of use (Williams and David, 1973; McLaughlin et al., 1996; Grant and Sheppard, 2009). Questions were raised above about the levels of Cd which are “potentially tolerable” in human diets because of different views on which dietary Cd levels caused the ﬁrst adverse effects on humans in a lifetime, the bioavailability of Cd in different foods, and the role of crop Zn inhibiting crop Cd absorption by humans. But confusion about the basis for limits has prevented the needed controls from being established in many jurisdictions. Clearly, some limit on Cd in phosphates is needed. Table 2 shows the present limits for a 45% P2O5 product (triplesuperphosphate) in different control schemes. Because some jurisdictions limit annual or 45-year cumulative applications, one must assume a P2O5 application rate to estimate the limit for Cd in the product, so in the present comparison, 98 Rufus L. Chaney

1 1 100 kg P2O5 ha yr was assumed (say for potatoes, much higher rate than for wheat). Based on the review above, it is appears that the EU 1 proposal of 20 mg Cd kg P2O5 may be unnecessarily low, and the 1 AAPFCO proposed 450 mg Cd kg P2O5 may be too high to limit crop Cd in the long term based on the Australian experience. The co- contamination with Zn reduces risk from Cd in some products, but this is not considered in present proposals or rules. Some phosphate ore reserves will require Cd removal before products should be marketed. But Cd-free products are not required to protect public health over centuries of risk consideration, perhaps longer than known P reserves are predicted to last (note above discussion of lack of increase in crop Cd level over time despite continuing use of P-fertilizers and other Cd sources for cropland).

10. Need for Regulatory Enforcement on Composition of Soil Amendments

The contaminated ZnSO4 fertilizer case discussed above is a valuable example of why regulatory controls are needed. In the USA, States regulate fertilizer products; initially to assure adequate levels of the purchased nutrients, and later to assure lack of contamination. Representatives of the 50 States work thru the AAPFCO to develop the basis for uniform regulations which would be enforced at the State level. Uniform regulations support interstate commerce and are necessary in a large nation. In most US States, the Depart- ment of Agricultural has staff who actually sample products around their state which is the reason companies need to comply with the rules. Companies have to analyze potential products and submit data to obtain a license to sell fertilizer products, and they have to analyze their products often and report at least annually to the State governments where they market. Products must comply with the State rules or they may not be marketed. Other than Cd in phosphates, few commercial fertilizer products have ever exceeded regulated levels of trace elements. But some garden and lawn fertilizers have contained very high levels of Pb and As and some other contaminants. Fortunately when the product “Ironite” was initially marketed with 4000 mg kg1 of As and Pb, States could apply their rules. Some states prohibited sale while other States required a package listed application rates to low that the product would not provide any beneﬁt. When they were unable to obtain permits to market their product, the company found a different source of mining byproduct which did not have such high contamination with Pb and As or other unwanted materials. Contrast this with the international sale of the Cd-contaminated Zn fertilizer product described above. Some nations had essentially no Food Safety Issues for Mineral and Organic Fertilizers 99 impediment to open marketing of the product. And other nations with good rules did not have the ability to require pre-analysis and permit issuance before the product could be sold for use as a fertilizer or feed ingredient e Even in the EU and Australia. Control or regulatory programs are needed to limit this kind of adulterated fertilizer product from being marketed anywhere, but especially in lands were subsistence rice farming is common. Effective rule enforcement can prevent any of the risks discussed in this paper from happening in practice.

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