Oral Bioavailability of As, Pb, and Cd in Contaminated Soils, Dust, and

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Oral Bioavailability of As, Pb, and Cd in Contaminated Soils, Dust, and Foods based on Animal Bioassays: a Review Hong-bo Li, Meng-Ya Li, Di Zhao, Jie Li, Shiwei Li, Albert L. Juhasz, Nicholas T. Basta, Yongming Luo, and Lena Q. Ma Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.9b03567 • Publication Date (Web): 23 Aug 2019 Downloaded from pubs.acs.org on August 24, 2019

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Environmental Science & Technology

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Oral Bioavailability of As, Pb, and Cd in Contaminated Soils, Dust, and Foods based on

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Animal Bioassays: a Review

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Hong-Bo Li,†,* Meng-Ya Li,† Di Zhao,† Jie Li,‡ Shi-Wei Li,‖ Albert L. Juhasz,§ Nicholas T.

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Basta††, Yong-Ming Luo‡‡, Lena Q. Ma†,┴,*

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†State

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Nanjing University, Nanjing 210023, China

Key Laboratory of Pollution Control and Resource Reuse, School of the Environment,

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‡College

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‖School

of Water Conservancy and Environment, University of Jinan, Jinan 250022, China

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§Future

Industries Institute, University of South Australia, Mawson Lakes, South Australia

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5095, Australia

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††

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43210

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‡‡

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Chinese Academy of Sciences, Nanjing 210008, China

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┴Soil

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United States

of Geography and Environment, Shandong Normal University, Jinan 250358, China

School of Environment and Natural Resources, Ohio State University, Columbus, Ohio

Key Laboratory of Soil Environment and Pollution Remediation, Institute of Soil Science,

and Water Science Department, University of Florida, Gainesville, Florida 32611,

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*Corresponding

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School of the Environment, Nanjing University, Nanjing 210023, China; Tel./fax: +86 025

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8968 0637, E-mails: [email protected]; [email protected]

author, State Key Laboratory of Pollution Control and Resource Reuse,

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ACS Paragon Plus Environment


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ABSTRACT. Metal contamination in soil, dust, and food matrices impacts the health of

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millions of people worldwide. During the past decades, various animal bioassays have been

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developed to determine relative bioavailability (RBA) of As, Pb, and Cd in contaminated

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soils, dust, and foods, which vary in operational approaches. This review discusses the

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strengths and weaknesses of different animal models (swine and mice), dosing schemes

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(single gavage dose, repeated gavage dose, daily repeated feeding, and free access to diet),

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and endpoints (blood, urine, and tissue) in metal-RBA measurement; compares metal-RBA

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obtained using mouse and swine bioassays, different dosing schemes, and different

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endpoints; and summarizes key findings on As-, Pb-, and Cd-RBA values in contaminated

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soils, dust, and foods. Future directions related to metal-RBA research are highlighted,

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including 1) comparison of metal-RBA determinations between different bioassays and

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different laboratories to ensure robust bioavailability data, 2) enhancing the metal-RBA

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database for contaminated dust and foods, 3) identification of physiological and

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physicochemical mechanisms responsible for variability in metal-RBA values, 4) formulation

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of strategies to decrease metal-RBA values in contaminated soils, dust, and foods, and 5)

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assessing the impacts of co-contaminants on metal-RBA measurement.

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TABLE OF CONTENT

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Introduction Industrialization and urbanization have contributed to metal contamination in the

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environment, adversely impacting human health.1 Among metals, arsenic (As), lead (Pb), and

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cadmium (Cd) are the most-studied and have been linked to various diseases including

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cancers.2-4 Among exposure pathways, incidental ingestion of soil/dust is an important

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contributor to metal intake, especially for children.5 A median soil ingestion rate of 52 mg/d

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was estimated for the general population of Chinese children aged 2.5–12 years,6 while the

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rate was 51–94 mg/d for children aged 0.5–9 years at a contaminated site in the US.7 Data

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showed association between childhood blood Pb level and Pb concentrations in contaminated

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soils and dusts,8-9 which was further supported by a close agreement in stable isotopic

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composition between childhood blood Pb and Pb contents in contaminated soils/dusts.5,10-11

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In addition to non-dietary soil/dust ingestion, daily consumption of food (rice, wheat,

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and vegetables) is another important contributor to metal exposure.12-14 Metals in soils,

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especially As and Cd, can be taken up by crops and accumulate in cereal grain and edible

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parts of vegetables.15-17 This is especially true for rice, which contributes ~70% of food intake

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for Asian populations.18-19 Elevated As and Cd concentrations have been reported in rice,

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which result from both geogenic and anthropogenic sources.20-22

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To accurately assess human exposure to metals in contaminated soil, dust, and food

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matrices, measurement of their metal bioavailability is necessay.23-26 The difficulty in

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measuring bioavailability in humans has inspired the development of surrogate animal

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bioassays.26 Using animals, bioavailability can be measured as absolute bioavailability, which

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is the fraction of ingested metal that is absorbed and reaches the systemic circulation, i.e., the

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ratio of absorbed dose versus administered dose.27 However, due to physiological differences,

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metal absorption between animals and humans may vary, thereby limiting the extrapolation

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of animal absolute bioavailability data to human health risk assessment. To overcome this

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drawback, oral bioavailability is often measured as relative bioavailability (RBA), which

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compares metal accumulation in animal tissue or urine following exposure to test materials to

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that of a soluble reference such as sodium arsenate (NaH2AsO4), lead acetate (Pb(AC)2), and

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cadmium chloride (CdCl2).28-30 In this case, the differences in metal absorption between

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animals and humans can be largely reduced. As such, metal-RBA measured using animal

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bioassays can be used for human health risk assessment.27

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To measure metal-RBA in soil, dust, and food samples, different bioassays have been 4


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developed, with mouse and swine models being most common.26-28 In addition, different

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dosing schemes and bioavailability endpoints have been used in these bioassays.

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Understanding the strengths and weaknesses of bioassay methodologies helps researchers to

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select a suitable animal model to accurately determine metal-RBA in target materials. While

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contaminated soils has been the focus of bioavailability studies,29-40 recent studies have

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expanded metal-RBA research to contaminated dust and food matrices.5,41-45 Thus, an

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overview of RBA in contaminated soils, dust, and foods provides researchers the typical

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RBA values in different matrices.

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This review assesses recent advances in animal bioassay methodologies and

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summarizes As-, Pb-, and Cd-RBA in contaminated soil, dust, and food matrices. Specific

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objectives were to: (1) discuss the strengths and weaknesses of different animal models,

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dosing schemes, and endpoints in metal-RBA measurement; (2) compare metal-RBA

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obtained using mouse and swine bioassays, different dosing schemes, and different

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endpoints; and (3) summarize As-, Pb-, and Cd-RBA values from different matrices using

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mouse and swine bioassays. In addition, future directions related to metal-RBA research are

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highlighted.

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Animal Models to Measure Metal Bioavailability

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The most common animal models for metal-RBA studies are swine and mouse,

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although rat, dog, monkey, and rabbit have also been used.46-52 Table 1 summarizes the

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strengths and weaknesses of the two animal models. Swine share morphological and

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physiological similarities with children (gastrointestinal tract, physiologic age, body weight,

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bone density, and mineral metabolism), thereby serving as an attractive model for

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bioavailability assessment.53,54 In addition, swine have large body size, allowing repeat blood

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sampling over exposure periods, thus reducing the numbers of animals required for RBA

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assessment.53,54 However, swine are not easy to handle in laboratories as it often requires

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collaborations with large animal facilities for RBA studies. Another limitation is that

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depending on bioavailability endpoint, surgical procedures may be required for the insertion

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of jugular catheters for repeat blood sampling, which requires specialized facilities and

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expertise.

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Mice are widely used in medical research due to their well-characterized physiology,

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genome, and potential to be manipulated such as altered genotype and dietary components.54

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Mouse models are a cornerstone of modern medical research to identify and test new drugs 5



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and investigate the associated pathophysiological mechanisms.55 Although lineages are

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entirely different, more than 85% of genomic sequences are conserved between mice and

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humans.56 In addition, mice share some similarities with humans in anatomical, histological,

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and physiological features in the intestines.57 However, there are differences in metal

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metabolism. For example, following ingestion, inorganic As undergoes biomethylation,

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which involves reduction of arsenate (AsV) to arsenite (AsIII) and oxidative methylation. In

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humans, biomethylation stops at dimethyl arsenicals, with inorganic As, monomethyl arsenic

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(MMA) and dimethylarsinic (DMA) being detected in the urine.58-59 However, in mice,

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biomethylation processes end at the trimethyl arsenicals, mainly trimethylarsine oxide.60

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Despite the differences in As methylation, the similarities are sufficient to use mice as a

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model to measure metal-RBA in contaminated matrices for human health risk assessment.61-62

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Compared to swine, mice are cost-effective and thus have the potential to be used in large

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sample sizes. In addition, mice are easily bred in laboratories as there is no need for surgery

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to collect blood and tissue, although repeat blood sampling is not possible.

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However, the physiological differences between swine and mice might bring

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differences to RBA measurements. Swine have a larger stomach, allowing significantly

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higher stomach fluid content (~1500 mL) than mice (0.60 for

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mice, suggesting that >60% of ingested inorganic As is absorbed and excreted in the

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urine.43,68 As such, urinary As analysis is a sensitive measure for As-RBA determination,

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especially at low-As doses.5,43 Low-As doses often produce unmeasurable As accumulation

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in tissue but measurable As in the urine. However, it should be noted that urine collection

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from swine or mice is prone to contamination by feces (Table 1). To overcome this,

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metabolic cages are used to separate urine and feces for mice. For swine, catherization has

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been used, but this requires a surgical procedure, which may cause animal infection.

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The daily repeated feeding (DRF) scheme has been widely used to determine

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As-,31,79-87 Pb-,31,88-91 and Cd-RBA31,92 in contaminated soils using a swine model. By using

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the DRF scheme, the disadvantages of gavage dosing can be overcome, i.e., sampling

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complexity for swine, multiple sacrifice for mice, and injury associated with gavage.

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However, the DRF scheme requires longer exposure timeframes to measure metal-RBA, i.e.,

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10–15 d compared to 26–48 h for the SGD scheme. With longer exposure timeframes, this

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may lead to an advantage of improved metal-RBA repeatability between replicate animals

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(Table 1). Due to the physiological differences between individual animals, the SGD scheme

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often produces higher variability in blood metal concentrations compared to metal

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accumulation in tissue following longer exposure timeframes. For example, the average

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relative standard deviation (RSD) of As-RBA values in 25 contaminated soils by

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swine-SGD-blood AUC was 23.9 ± 18.7%,28,32 ~2 times higher than that of 14.0 ± 8.46% for

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24 soils by swine-DRF-urine bioassay.87

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Free Access to Diet (FAD). For mice, a similar approach to the swine-DRF scheme

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has been developed, i.e., free access to diet (FAD) with tissue metal or urinary As as the

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bioavailability endpoint (Table 1). Bradham et al.93-94 determined As-RBA in contaminated

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soils using this bioassay. Prior to exposure, soil-amended diets were prepared by mixing soil

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with powdered rodent diet (1% w/w). Mice (female C57BL/6 mice) housed in individual 10


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cages (3 mice in a cage used as an operation unit) had free access to soil-amended diets for 9

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d, which was followed by a 1-d clearance period. During exposure, daily urine samples were

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collected and measured for volume and As concentration. The mouse-FAD-urine bioassay

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has been used to measure As-RBA in contaminated soil and rice samples.5,30,37,43

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Besides As, the mouse-FAD-tissue bioassay has also been utilized to determine Pb-

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and Cd-RBA in contaminated soils, dust, and foods. For example, Juhasz et al.29 used this

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bioassay to determine Cd-RBA in contaminated soils with Cd accumulation in the liver

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and/or kidneys as the bioavailability endpoint. The experiment was conducted with diets

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amended with soil or CdCl2 over 15 d using male Balb/c mice. The Cd concentrations in the

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tissue following FAD dosing were dose-normalized to calculate Cd-RBA in soil (Eq. 2). The

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mouse-FAD-tissue bioassay is preferred for Pb- and Cd-RBA measurement because they

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both mainly accumulate in the tissues (liver, kidneys, and/or bone) following absorption.

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With high sensitivity, this bioassay can also be used to determine Pb- and Cd-RBA at

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low-dose concentrations for contaminated rice, wheat, and vegetable samples.5,44-45

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Compared to the other three dosing schemes (SGD, RGD, and DRF), the FAD

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approach supplies mice with soil- or food-amended diets, which are easy to prepare and

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operate. Following preparation, diets are supplied to mice for free consumption, which is less

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stressful to mice compared to gavage. In addition, the FAD scheme more closely mimics the

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status of children’s daily exposure to soil/dust ingestion or daily food consumption (Table 1).

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Comparison between Different Bioassays

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Animal bioassays have been developed to determine metal-RBA in environmental

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matrices, which include different animal models (mouse and swine), dosing schemes (SGD,

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RGD, DRF, and FAD), and bioavailability endpoints (metals in blood, urine, and tissue)

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(Table 1). Since these bioassays differ in physiological parameters, it may influence metal

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absorption in the gastrointestinal tract, thereby influencing metal-RBA measurement. To

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date, there are limited reports comparing metal-RBA between different bioassays. Such

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information may help reduce regulatory uncertainty as to the most appropriate assay for

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measuring metal-RBA for the refinement of exposure for human health risk assessment.

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Swine vs. Mouse Model. Both swine and mouse models have been widely used to

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measure metal-RBA in contaminated soils, but few studies have compared the animal models

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basing on the same dosing scheme. Using different dosing schemes, Bradham et al.68 11



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compared As-RBA in 9 soils and 3 standard reference materials using swine-DRF and

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mouse-FAD bioassays with urinary As as the endpoint. A linear correlation (r2=0.49)

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between the two bioassays was observed, with the swine bioassay producing higher As-RBA

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for 9 soils (mouse/swine As-RBA ratios = 0.4–0.8), while for the 3 standard reference

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materials, the difference in As-RBA was insignificant (mouse/swine ratios = 0.9–1.0). Also,

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using different dosing schemes, Bradham et al.95 measured As-RBA in 20 contaminated soils

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using the mouse-FAD-urine bioassay, while 7 of the 20 soils were assessed using the

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swine-DRF-urine bioassay with the remaining 13 soils assessed using the swine-SGD-blood

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bioassay. Similarly, the swine and mouse bioassays showed a linear relationship with the

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swine bioassay providing higher As-RBA (As-RBAswine = 1.23 × As-RBAmouse – 0.80,

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R2=0.50). Basta et al.96 and Stevens et al.97 reported As-RBA for 14 contaminated-soils using

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both the mouse-FAD-urine bioassay (mean As-RBA = 26.4%) and the swine-DRF-urine

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bioassay (mean As-RBA = 36.5%). A linear relationship was found between As-RBA

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determined using swine and mice, again with swine providing higher As-RBA (As-RBAswine

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= 0.96 × As-RBAmouse + 11.3, R2=0.65). The distribution was bimodal with 7 soils exhibiting

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higher As-RBA for swine compared to mice but for the remaining 7 soils, As-RBA was

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equivalent between the two animal models. However, these evaluations are not direct

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comparisons between swine and mouse models as the studies differed in dosing schemes.

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Higher As-RBA using the swine bioassay may be partially due to use of SGD or DRF

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schemes, which measure As-RBA under a fasted state or limited feeding conditions. This

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may enhance metal dissolution from soil compared to the FAD scheme due to differences in

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gastric phase pH.

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Up to now, the only true comparison of the swine versus mouse model based on the

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same dosing scheme and endpoint was by Li et al.37 who compared As-RBA in 12

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contaminated soils measured using the mouse-SGD-blood and swine-SGD-blood bioassays.

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Although animal models varied, the dosing scheme and bioavailability endpoints were

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similar. Li et al.37 showed that there was no significant difference in As-RBA between the

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two animal models for 8 of the 12 soils, with the two data sets showing a strong linear

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correlation (As-RBAswine-SGD-blood = 1.04× As-RBAmouse-SGD-blood +8.10, R2=0.83). The data

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suggest that when using the same dosing scheme and endpoint, As-RBA is consistent

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between swine and mouse models. However, compared to swine, the mouse model offers the

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advantages of lower cost and easier care, hence having the potential to be used to assess large

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sample sizes. 12


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Unlike As, studies have not compared Pb- and Cd-RBA measurements using different

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animal models with the same dosing scheme and bioavailability endpoint. However,

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Bradham et al.69 measured Pb-RBA in SRM 2710a (Montana Soil 1 from NIST; 49%) by

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combining different bioavailability endpoints (blood, liver, kidneys, and bone) using the

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mouse-FAD bioassay. Their data were comparable to Pb-RBA (57%) determined using the

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swine-DRF bioassay by Casteel et al.98 In addition, Pb-RBA measured using swine-DRF and

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mouse-FAD models in an untreated and a phosphoric acid-treated soil sample was assessed

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by Bradham et al,99 showing that based on 90% confidence limits, Pb-RBA based on

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mouse-bone bioassay was higher than corresponding swine RBA estimates. To ascertain

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whether the mouse model is a useful alternative to swine for Pb- and Cd-RBA measurement,

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comparison of Pb- and Cd-RBA for a larger number of samples is warranted.

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Different Dosing Schemes. Besides animal models, it is also important to compare

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different dosing schemes for a given animal model. In addition to mouse-SGD-blood

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bioassay, Li et al.37 also measured As-RBA in 12 contaminated-soils using the mouse-FAD

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bioassay, with urinary As and tissue As as the endpoints. Based on urinary As, there was no

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significant difference in As-RBA when the FAD and SGD schemes were compared for 9 out

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of 12 soils. Further, As-RBA determined using the FAD-urine bioassay was correlated with

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data using the SGD-blood bioassay (As-RBAFAD-urine = 0.86 × As-RBASGD-blood +7.33,

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R2=0.87). Similarly, strong linear correlations were found between As-RBA determined

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using SGD-blood and FAD-liver bioassay (As-RBAFAD-liver = 0.97× As-RBASGD-blood+6.19,

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R2=0.88) or FAD-kidney bioassay (As-RBAFAD-kidney = 0.65× As-RBASGD-blood +7.54,

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R2=0.66), with no significant difference in the slopes of the 3 correlations.

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The data suggest that for the mouse model, different dosing schemes have little effect

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on As-RBA measurement. Compared to the SGD-blood bioassay, which require large

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numbers of mice to establish As-AUC, the FAD-tissue bioassay has the advantage of using

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small animal numbers. For a given soil sample, 18 mice are required to collect blood samples

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from triplicate mice at 6 time-intervals for As-AUC, while only 3 mice are needed using the

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FAD-tissue bioassay to determine As-RBA.

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Besides As, using a mouse model with two different dosing schemes and endpoints,

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i.e., SGD-blood, and FAD-tissue, Li et al.40 also measured Pb-RBA in 12 contaminated-soils.

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Results showed no significant difference in Pb-RBA between the two dosing schemes.

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However, compared to the SGD-blood bioassay, which often shows large RSD among 13



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replicates (averaging 35%), the FAD-tissue bioassay shows lower RSD (averaging 11%),

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suggesting that the FAD-tissue bioassay is more robust in determining Pb-RBA in soils.

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Other studies also showed RSD of ~30% for Pb-RBA using the SGD-blood bioassay

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compared to that of 3.2–5.1% for the FAD-tissue bioassay.30,31,34 This is because multiple

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doses of soil via daily diet consumption over a period (e.g., 10 d) reduces the variation in Pb

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dissolution and absorption in the gastrointestinal tract among animals, thereby lowering

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variation in Pb-RBA measurement.

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Different Endpoints. Depending on animal model, dosing scheme, and target metals,

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different endpoints may be selected (blood, urine, and/or tissue) for RBA measurement

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(Table 1). For the SGD scheme, blood is often used,41-42 while urine and tissue are used for

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the RGD, DRF, and FAD schemes.5,36-40,43-45,77 Among metals, As is the only one that uses

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urine as the endpoint.93 For As, all 3 endpoints may be used for RBA measurement,37 while

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blood and tissue are suitable endpoints for Pb.34,36,41 However, only tissue accumulation has

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been used as an endpoint for Cd-RBA measurement.29,38

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Limited studies have compared metal-RBA measurements using different endpoints.

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Using the mouse-FAD bioassay, Li et al.37 compared As-RBA determined using urinary As

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and tissue accumulation as the endpoint. When As-RBA values in 12 contaminated-soils

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were compared, values determined using liver and kidney accumulation showed a strong

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linear correlation (As-RBAkidney = 0.65 × As-RBAliver + 4.98, R2=0.75), with kidney showing

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significantly lower As-RBA for 4 soils. In addition, As-RBA determined using liver and

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kidney was similar to those determined using urinary As (As-RBAliver = 1.07 × As-RBAurine –

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1.11, R2 = 0.89; As-RBAkidney = 0.82× As-RBAurine + 0.29, R2 = 0.88) with 2–3 soils showing

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significant difference in RBA estimations. The data suggest overall consistent values were

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obtained using different endpoints to measure As-RBA in soils using the mouse-FAD

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bioassay.

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Based on the swine-DRF bioassay and different endpoints, Denys et al.31 also showed

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agreement between As-RBA based on As accumulation in the kidneys and liver (R2=0.98),

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between As in the liver and urinary As (R2 = 0.99), and between As in the kidneys and

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urinary As (R2 =0.99). Compared to determination of metal concentrations in the liver and

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kidneys, precise urine collection and metal quantification is challenging due to the potential

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for fecal contamination when using metabolic cages. Based on Li et al.37, the mouse-FAD

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bioassay using liver or kidney accumulation as the endpoint is robust, easy to handle, which 14


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is recommended for metal-RBA analysis, especially for large sample sizes.

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Metal Relative Bioavailability in Soil, Dust, and Food Samples

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Using different animal bioassays, As-, Pb-, and Cd-RBA in contaminated soils, dust,

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and foods has been determined. Tables 2–3 summarize key studies using swine and mouse

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models. Some early reports of metal-RBA in contaminated soils using dogs46, rabbits,47

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monkeys,48-50 and rats are excluded since these animal models are rarely used.51-52

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As-, Pb-, and Cd-RBA Values in Contaminated Soils. Up to now, most RBA

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studies have focused on contaminated soils (Table 2A-B). Tables S1–S3 summarize As-,

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Pb-, and Cd-RBA in contaminated soils together with soil properties (total As, Pb, Cd, Fe, Al,

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and P, and soil pH). Eighteen studies have utilized swine (SGD or DRF scheme) to measure

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As-RBA in contaminated soils, while 6 studies have used mice to measure As-RBA (SGD,

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FAD, or RGD schemes), with As in blood, urine, or tissue as endpoints (234 contaminated

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soils; 18–25,000 mg kg–1 As) (Table 2A). For these soils, As-RBA showed considerable

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variation (1.3–100%) with mean and median values of 30.2±18.5% and 27.4%, respectively

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(Figure 1A). Most contaminated soils had As-RBA value lower than the default value of

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60% proposed by USEPA.23 Similar to As, Pb-RBA has been measured in 155 contaminated

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soils at 1.0–140% (12.6–40,214 mg kg–1 Pb), with mean and median values of 49.5±28.2%

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and 50.8%, respectively (Table 2B). However, to our knowledge, only 4 studies have been

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performed using swine or mice to assess Cd-RBA (39 contaminated soils; 3.0–465 mg kg–1

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Cd) (Table 2B), with values ranging from 10–116% (mean and media values of 55.1±25.4%

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and 56.2%, respectively) (Figure 1A).

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For As- and Pb-RBA, there are decreasing trends with increasing total As, Pb, and Fe

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concentrations in soils, while they generally increase with increasing soil pH (Figure 2A-F).

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However, there seems no relationship between Cd-RBA and total Cd, while Cd-RBA tends to

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decrease with increasing Zn/Cd molar ratios (Figure 2H-J). The decrease in As- and

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Pb-RBA with increasing As and Pb concentrations could be explained by the fact that

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increasing amounts of As and Pb in soils may be present as insoluble species when soil is

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grossly contaminated from mining activities. In soils, metals are often adsorbed by Fe oxides.

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With increasing soil Fe concentration, stronger binding of metals by Fe oxides may limit As

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and Pb dissolution from soils in gastrointestinal fluids, thereby decreasing As-and Pb-RBA.

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In the gastrointestinal tract, Pb and Cd are absorbed via divalent metal transporters, which are

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shared by mineral nutrients including Zn, Fe, and Ca.100 Due to competitive uptake, less Pb 15



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and Cd are absorbed with increasing mineral nutrition in animals.101-102 By understanding the

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mechanisms responsible for different As- and Pb-RBA with soil properties, mitigation

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strategies to reduce metal-RBA in soils can be developed.

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As-, Pb-, and Cd-RBA Values in Contaminated Dust. In addition to soils,

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incidental ingestion of dust is also an important exposure source for children. Deposition of

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fine air particulates in addition to tracking of outdoor soil inside, are predominant sources of

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indoor dust. Dust exposure from atmospheric deposition may be more significant in China

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and India due to a high frequency of haze episodes and high concentrations of fine airborne

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particulates.103 For example, a Chinese coastal city with relatively clean air had a housedust

431

loading rate of 13.1 mg m–2 d–1,104 significantly higher than those for the US (1.8–4.9 mg m–2

432

d–1), Germany (5.41 mg m–2 d–1), and Canada (11.1 mg m–2 d–1).105-107 In addition, families in

433

China mainly live in apartments without backyards, limiting children’s exposure to soil but

434

increasing the chance of housedust ingestion.

435

However, compared to contaminated soils, there are limited data on metal-RBA in

436

housedust samples (Tables 3 and S4). Collecting housedust samples (n = 12; 7.0–38.2 mg

437

kg–1 As and 63–738 mg kg–1 Pb) from urban cities and using the mouse-SGD-blood bioassay,

438

Li et al.41-42 reported As- and Pb-RBA at 22–86% (mean 51.0±17.9%) and 29–60% (mean

439

49.6±9.47%), respectively. Extending this research to contaminated housedust (37.2–1051

440

mg kg–1 As, 138–56,185 mg kg–1 Pb, and 3.70–329 mg kg–1 Cd) from mining/smelting

441

locations, Zhao et al.5 determined As-, Pb-, and Cd-RBA using the mouse-FAD bioassay with

442

liver and kidney accumulation as the endpoint (Table S4). Compared to As- and Pb-RBA for

443

urban housedust, housedust from mining/smelting locations showed lower As-RBA (8.5–

444

37%; mean = 16.7±6.71%) and Pb-RBA (11–34%; mean = 25.4±6.78%), suggesting the

445

influence of dust properties and contamination sources on metal-RBA. In addition, Zhao et

446

al.5 reported that Cd-RBA in housedust from mining/smelting locations was 28–68% (mean

447

45.6±9.98%), which was significantly higher than As- and Pb-RBA in the same samples

448

(Figure 1B), consistent with the pattern of metal-RBA in contaminated soils (Figure 1A).

449

As-, Pb-, and Cd-RBA Values in Contaminated Food. Similar to housedust, limited

450

data are available for metal-RBA in foods (Tables 3 and S5). Juhasz et al.72 was the first to

451

determine As absolute bioavailability in rice using the swine-SGD-blood bioassay. The As

452

absolute bioavailability in rice dominated by inorganic As was significantly higher than those

453

containing predominantly organic As (89 vs. 33%). This was due to differences in the 16


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bioavailability of inorganic As (AsIII: 104%; AsV: 93%) versus organic As (MMA: 17%;

455

DMA: 33%). Expanding this research, Li et al.43 and Zhao et al.5 determined As-RBA in rice

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samples from Chinese markets (0.13–0.32 mg kg–1 As; n = 11) and mining/smelting areas

457

(0.17–0.49 mg kg–1 As; n = 11) (Table S5). Based on the mouse-FAD-urine bioassay,

458

As-RBA was 45–88% (67±13%) and 11–65% (31±12%), respectively. For rice from

459

mining/smelting areas, As-RBA showed a positive correlation with inorganic As while

460

As-RBA was negatively correlated with DMA, suggesting As speciation affects As-RBA in

461

rice.5

462

Besides As, Cd in foods, particularly rice, is also of health concern. While attention

463

has been paid to Cd concentrations in rice, limited studies have measured Cd bioavailability.

464

Zhao et al.45 collected food samples including rice, wheat, and vegetables and human urine

465

samples from a Cd-contaminated area in Yixing, China, which is known for enamel pottery

466

production. Using the mouse-FAD-kidney bioassay, Cd-RBA in rice (n=10, 0.29–1.09 mg

467

kg–1), wheat (n=8, 0.70–1.98 mg kg–1), and vegetables (n=6, 0.35–2.33 mg kg–1) was

468

measured at 17–57 (42±12%), 37–68 (48±9.3%), and 18–78% (49±21%), respectively. When

469

urinary Cd concentrations from local residents were predicted from Cd intake based on total

470

Cd in rice, predicted values (4.14 g g–1 creatinine) overestimated measured urinary Cd (1.20

471

g g–1). However, when urinary Cd was predicted by incorporating Cd-RBA in rice (1.07 g

472

g–1), a close agreement with measured values was obtained.45

473

In addition, Zhao et al.5 determined Cd-RBA in 11 rice samples (0.41–1.67 mg kg–1)

474

from mining/smelting areas in Hunan, China with values at 41–84% (59±13%), higher than

475

those in Yixing from Zhao et al.45 (17–57%; 42±12%). The mechanisms controlling variation

476

in Cd-RBA in rice are still unclear. As there are no specific Cd transporters, Cd is absorbed

477

via transporters responsible for absorption of essential nutrients including Ca, Fe, and Zn.100

478

It has been shown that mineral nutrients play an important role in decreasing Cd absorption

479

from rice via competition for shared transporters. This has been supported by the fact that

480

dietary amendments including Ca, Fe, and Zn effectively reduce Cd-RBA in rice.44 The

481

difference in mineral nutrients (Ca, Fe, and Zn) in rice from different production areas may

482

explain differences in Cd-RBA values, although further studies focusing on associated

483

mechanisms are required.

484 485

Compared to As and Cd, only one study measured Pb-RBA in rice using a mouse model (Tables 3 and S5). By performing the mouse-FAD-tissue bioassay on samples from 17



486

mining/smelting sites, 9 rice samples with Pb concentrations of 0.30–0.88 mg kg–1 were

487

measured for Pb-RBA, with values ranging from 11–59% (mean = 31±15%).5 Pb-RBA was

488

lower than As-RBA (50±22%, n=22) and Cd-RBA (51±15%, n=21) (Figure 1C), suggesting

489

that Pb in food matrices is poorly absorbed.

490

Future Research Directions

491

Over past decades, advances have been made in the development and application of

492

animal models for determining metal-RBA in environmental matrices for application to

493

human health exposure and risk assessment. Without considering RBA, the underlying

494

assumption in quantifying metal intake from soil, dust and food matrices for human health

495

risk assessment is that all contaminants measured by total analysis is equal to the absorbed

496

dose. However, for an adverse health effect, the metal must be dissolved in the

497

gastrointestinal tract for absorption to occur across the intestinal epithelium into the systemic

498

circulation. Therefore, measuring metal-RBA is important for adjusting the concentration

499

term to account for differences between conditions under which the toxicity criterion is

500

derived and conditions being quantified in site-specific risk assessment.

501

Metal-RBA describes the absorbed fraction of a chemical from a particular exposure

502

medium relative to the fraction absorbed from the dosing vehicle used in the toxicity study.

503

Such data are often lacking for environmental matrices but can be measured using robust in

504

vivo animal bioassays. Understanding exposure is critical as outcomes from human health

505

risk assessment may have social, economic, and environmental impacts through risk

506

management and remediation decisions. Although standard operating procedures are

507

available for swine and mouse bioassays, future research on metal-RBA is essential to

508

advance the science of metal exposure and risk assessment.

509

Importance of Animal Models, Dosing Schemes, and Endpoints. Although studies

510

have shown close agreement between As-RBA measurements between different animal

511

models, dosing schemes, and endpoints for contaminated soils,37,68,95 limited data are

512

available for other As-impacted matrices and for Pb- and Cd-RBA measurements. Comparing

513

the influence of animal model and bioassay operational parameters on metal-RBA from soil,

514

dust, and food matrices is critical to understand the impact of RBA methodology on exposure

515

refinement.

516

Inter-Laboratory Studies. To date, inter-laboratory studies have not been 18


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517

undertaken to assess the robustness and transferability of in vivo bioassays from one

518

laboratory to another. Understanding inter-laboratory variability is important for the

519

development and application of robust bioassays for use worldwide. Initially, intra-laboratory

520

studies should be undertaken to assess repeatability and reproducibility at the quantification

521

range and at regulatory levels. The magnitude of random errors may be determined through

522

repeatability and reproducibility assessment following assessment at different times and with

523

different operators. Inter-laboratory studies should determine the robustness and

524

transferability of the in vivo bioassays including replication, blind replication in addition to

525

procedural blanks and spiked solutions. This is not a trivial undertaking as in vivo bioassays

526

are costly and require specialized skills and facilities, particularly for swine bioassays.

527

Limited Metal-RBA Data for Dust and Food Matrices. As detailed in Table 2, the

528

majority of metal-RBA data have been generated for contaminated soils as there are limited

529

data available for dust and food matrices (Table 3). Future studies should focus on generating

530

data on metal-RBA in dust and food to expand the existing data set. For dust, the majority of

531

data have been generated from samples collected from urban environments (low metal

532

concentrations), so a research focus should be on dust from residential areas close to

533

industrial activity. Although studies have shown that As speciation in rice strongly influences

534

As-RBA,5,43,72 the mechanisms accounting for the variability in Cd- and Pb-RBA in rice,

535

wheat, and vegetables are unclear.5,45 Understanding RBA drivers are important since this

536

information may be utilized to facilitate the development of mitigation strategies to decrease

537

metal-RBA in food.

538

Mechanisms Impacting Metal-RBA. Metal exposure from environmental matrices

539

is complicated due to interactions between metals, interactions with matrix adsorbent phases,

540

mineralogical constraints, and competitive uptake processes between non-essential and

541

essential elements. Studies assessing the influence of metal speciation on metal-RBA have

542

identified that speciation alone does not predict metal-RBA outcomes. Based on synchrotron

543

X-ray techniques, Bradham et al.93 observed lower As-RBA for soils containing As sulfides

544

(realgar or arsenopyrite), suggesting the presence of insoluble arsenopyrite or realgar reduced

545

As-RBA. Although significant (p < 0.10), the R2 value of 0.28 for the predictive equation

546

indicated that the overall fit was poor, supporting that As speciation alone does not accurately

547

predict As-RBA. Similarly, Stevens et al.97 reported As speciation was not predictive of

548

As-RBA for 27 contaminated-soils and solid-wastes. Principal component analysis found

549

67% of the As-RBA variation was explained by ferric arsenates, sorbed to amorphous ferric 19



550

arsenate, and co-preciptated with jarosite.97 Additional research is needed to assess other soil

551

properties influencing metal-RBA prediction. The lack of knowledge regarding the effects

552

and mechanisms of these interactions creates uncertainty regarding metal exposure, thereby

553

limiting the development and application of physicochemical and /or nutritional strategies to

554

minimize metal exposure.

555

Strategies for Reducing Metal-RBA. To better protect human health from metal

556

exposure via incidental soil/dust ingestion and food consumption, further research should be

557

conducted to formulate strategies to decrease metal-RBA. Although studies have shown the

558

effectiveness of phosphate amendments to decrease Pb-RBA in contaminated soils,76,99 the

559

efficacy is variable and depends on soil properties. The efficacy of other soil amendments

560

(e.g., Fe/Mn minerals) in reducing metal bioaccessibility have been assessed using in vitro

561

approaches, however, their effectiveness in vivo is yet to be determined. For example, Beak

562

et al.108 showed that Pb adsorbed to birnessite had low Pb bioaccessibility, suggesting that

563

birnessite may be effective in reducing Pb exposure from contaminated soils. Ferrihydrite

564

was also able to decrease As bioaccessibility (to

Oral Bioavailability of As, Pb, and Cd in Contaminated Soils, Dust, and

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