ARSENIC AND LEAD BIOACCESSIBILITIES MEASURED WITH A NOVEL REACTOR SYSTEM USING THE MEXICAN STANDARD AND PBET METHODS: COMPARISON WITH IN VIVO AND IN VITRO REPORTED DATA

BIOACCESIBILIDAD DE ARSÉNICO Y PLOMO DETERMINADA CON UN NOVEDOSO SISTEMA DE REACTORES USANDO EL MÉTODO PBET Y EL DE LA NORMA OFICIAL MEXICANA: COMPARACIÓN CON DATOS REPORTADOS IN VIVO E IN VITRO

The primary anthropogenic activities responsible for higher levels of potentially toxic elements (PTE) in the environment are energy production and mining-metallurgy (Panayotova 2016, Sposito 2008, Masindi and Muedi 2018). The biogeochemical cycles of several elements, including arsenic (As) and lead (Pb), have been disturbed (Masindi and Muedi 2018), and populations can be exposed to polluted soil and water. Depending on the dose and chemical species, As and Pb ingested may harm human health (Kumpiene et al. 2017). Other factors influence toxicity, mainly the route of exposure, solubility, particle size, environmental matrix type, and the presence of certain substances that reduce or improve their absorption. The Pb acetate is one of the most bioavailable compounds because it is very soluble (Freeman et al. 1994) but most of the Pb in soils is forming compounds with low solubility (Rooney 2002). In the gastrointestinal tract, iron, zinc, and calcium decrease Pb absorption, possibly competing for absorption receptors in the intestine. Biogeochemical transformations control As bioavailability, toxicity, and its environmental fate. As can be oxidated by diverse compounds or reduced by organic compounds, forming more toxic inorganic species. Direct reduction of Fe (III) by microorganisms can lead to As sequestration by sorption (Borch et al. 2010). These processes have a big influence on the As bioavailability in vivo and, consequently, on in vitro determinations.

Several toxic arsenic species may form in the intestine because of the reaction between arsenic and food compounds. (Conrad and Barton 1978, Mushak 1991, Diamond et al. 1998, Calatayud and Llopis 2015, Ollson et al. 2017)

The damages caused by arsenic are classified as carcinogenic, mutagenic, or genotoxic, producing immunological, reproductive, developmental, neurological, renal, hepatic, hematological, gastrointestinal, cardiovascular, pulmonary, respiratory, and dermal harm (Mandal and Suzuki 2002, Sattar et al. 2016). Pb is classified as “probably carcinogenic to humans” (Group 2A), according to IARC (2006). It has mainly neurocognitive and behavioral effects (Chiodo et al. 2004, Mason et al. 2014) and decreases children’s intellectual quotient (Schnaas et al. 2006, Mitra et al. 2017). However, because of various factors, most above mentioned, only part of the ingested element is absorbed through the gastrointestinal tract. This absorbed fraction is named bioavailable (Zhu et al. 2015). It can be measured as an absolute amount or a relative amount (Ruby et al. 1996, 1999, Juhasz et al. 2009, USEPA 2012, Koch et al. 2013).

Bioavailability is determined using animal models (in vivo), which is expensive, time-consuming (days or months), and raises ethical considerations. As an alternative to bioavailability tests, bioaccessibility tests mimic the human gastrointestinal processes (in vitro) being less costly and faster; these are also named in vitro bioavailability methods (Ruby et al. 1999, Juhasz et al. 2014, Liu et al. 2017).

Regulatory frameworks related to the remediation of polluted soils follow a risk assessment approach, and bioaccessibility is an adjustment factor in the calculations (USAEC, 2003). Mexican environmental regulation NOM-147-SEMARNAT/SSA1-2004 follows the same approach. Nevertheless, this regulation only allows the calculation of soil cleanup target levels based on gastric bioaccessibility of Pb measured through the simple bioaccessibility extraction test (SBET) (Dabin et al. 2012, Koch et al. 2013) using an end-over-end shaker.

In the Environmental Biogeochemistry Laboratory of the School of Chemistry of the National Autonomous University of Mexico (UNAM), PTE bioaccessibility is measured frequently. A novel agitation system called gastrointestinal simulation reactor system to determine bioaccessibility (GSRSB) was designed and built to facilitate this determination. It allows continuous control of the pH and improves gas flow stability, simplifying the reagents’ addition and aliquots sampling during phase change. Characteristics of the vessels of the agitation equipment avoids sedimentation and soil erosion (Quiroz-Vivanco 2018). This study aimed to use the GSRSB with the physiologically based extraction test (PBET) to measure Pb and As gastric and intestinal bioaccessibility in polluted soils, to show the performance of this novel equipment, and to have data to propose this method (GSRSB-PBET) as an alternative to the current NOM-147 standard method. As and Pb bioaccessibilities of three composite soil samples and the certified reference material from the National Institute of Standards and Technology (NIST) Montana High number 2710 (NIST 2710) were measured by GSRSB-PBET and NOM-147 methods to reach this goal, both methods results were compared. Additionally, only NIST 2710 results were compared with those reported in vivo and in vitro by 14 laboratories, applying 17 methods (Koch et al. 2013).

METHODS AND MATERIALS

Sample preparation

The polluted site was in a remediation process, and from the complete set of soil samples collected (N = 800), nine were selected based on their arsenic content, using a non-probability judgment sampling method (Frey 2018). The coordinates of the sites chosen are presented in table I. The nine samples (S1-S9) and wastes (converter flue dust, smelting furnaces flue dust, calcine, and black arsenic) were dried (40 ºC), ground, sieved (mesh #10 < 2 mm), and homogenized by quartering (Hesse 1971). Portions of 100 g were re-milled (ball grinder Fritsch), sieved (mesh #200 < 74 µm) (same particle size of NIST 2710 certified material) and preserved at room temperature (15-20 ºC) in hermetically sealed plastic containers. The three composite samples were prepared (C1, C2, and C3), mixing 100 g of three individual soils (Table I) selected by geographic proximity and As concentration: C1 (S1, S2, S3); C2 (S4, S5, S6) and C3 (S7, S8, S9).

TABLE I
SAMPLING SITE UTM COORDINATES, pH, AND TOTAL AND GEOAVAILABLE ELEMENT CONCENTRATIONS OF COMPOSITE SOIL SAMPLES, WASTES AND SUBPRODUCTS.

^a Zone 14; ^b Mean ± SD (n=3); ^c Detection Limit (D.L.) Pb =0.6 mg/kg, and Fe = 1 mg/kg; ^d ND: Not determined.

Analytical determinations

All chemicals used were analytical reagent grade (AR grade). Deionized water was used to prepare all solutions for the leaching tests and all analytical procedures. All determinations were done in triplicate, and their relative standard deviations (RSD) calculated (values in tables). All the extracts were preserved in the dark at 4 ºC. Elements were quantified using ICP-OES (Agilent Technologies model 5100), applying method 6010C (USEPA 2000). Digestions were done using an Ethos Easy microwave digestion system (Millestone Inc.) using Teflon PFA beakers applying US-EPA method 3051A (USEPA 2007b).

General composition and geoavailability

The pH in soils and wastes was measured following ISO-10390:2005 with a model Orion Star A211 Thermo Scientific potentiometer. Total concentrations of As, Pb, Fe, Ca, Cd, Cu, Zn, and Mn were measured with a portable model DP-6000-CC Thermo Scientific X-ray fluorescence Olympus analyzer using the 6200 method (USEPA 2007, Zamora-Martínez et al. 2008). Total sulfur was determined in a model S-832 Leco analyzer (Bremner and Tabatabai 1971). All composite soils were analyzed by SEM-EDS in a Hitachi TM1000 tabletop scanning electron microscope with an energy dispersive spectroscopy module, and by X-ray diffraction (XRD) with a Shimadzu XRD-6000 equipped with a Cu tube and a graphite monochromator. Blanks and certified reference materials were used for quality control for individual and composite soil analysis: pH (NIST-traceable buffer solutions), XRF (NIST 2711a), sulfur content (Oreas 45b) and ICP-OES determinations (HPS, QCS-26). Geoavailable concentrations in soils and wastes were determined with the D3987-06 method (ASTM 2006).

Soils bioaccessibility

Two methods have been used to measure the bioaccessibility of soils: NOM-147, which only includes the gastric phase and was applied as indicated by Mexican regulation (SEMARNAT 2007) using glycine at pH = 1.50 ± 0.05 (HCl) and an end-over-end shaker (37 ± 2 ºC). This method is analogous to the SBET, RBALP and SBRC (Dabin et al. 2012). The other method used was the PBET (Ruby et al. 1996), which determines gastric and intestinal phases. The reagents used include pepsin, citrate, malate, lactic acid, and acetic acid in the gastric phase and bile salts and pancreatin in the intestinal phase. Reagent quantities, ratios (solid/liquid), and residence times were applied as indicated by Ruby et al. (1996). A modification was made in the gastric phase, adjusting the pH to 1.50 ± 0.05 with HCl (11 M) as was reported to increased in vitro and in vivo correlations (Drexler and Brattin 2007, Juhasz et al. 2014). After the gastric phase concluded, pH was gradually increased to 7.00±0.10 with a NaHCO₃ saturated solution and subsequently, pancreatin and bile salts were added.

Intestinal extracts were digested (HNO₃-microwave-assisted, method 3051A) (USEPA 2007b) using an Ethos Easy (Milestone) to destroy organic compounds that interfere with PTE quantification by ICP-OES (USEPA 2000). Argon was used as a stirring gas in the novel reactor (GSRSB). As mentioned above, it was exclusively designed in our laboratory. It is used to improve the control pH and gas flow stability, simplifying the reagents’ addition and taking of aliquots during the phase change. It has conical bottom glass reactors placed in a heating bath (37±1 ºC). The vessels contain a thin tube through which the argon enters and stirs the suspension (1 L/min) (Fig. 1). The conical design is used to prevent the sedimentation and erosion of soil particles. The pH electrodes are placed inside each reactor, allowing measurements without stopping the agitation process. A small orifice is used to take aliquots or add reagents during the gastric-to-intestinal phase change and to introduce pH electrodes (García-Rodríguez 2017, Quiroz-Vivanco 2018). The NOM-147 end-over-end shaker (Fig. 1) controls the temperature with an immersion recirculation heater. Nevertheless, it does not allow one to measure pH, take aliquots, or add reagents when the shaker is in motion.

Fig. 1
Equipment used to measure oral bioaccessibility: (1A) End-over-end shaker NOM-147 method (gastric phase) and (1B) GSRSB-PBET method (gastric and intestinal phases).

In this study, the absolute bioaccessibility was calculated as the ratio between the element concentration (mg/kg) in the solution for each method or extraction phase, respect to the element concentration in the soil (mg/kg) in percentage. To calculate the relative bioaccessibility, the bioaccessibility results were corrected with the mean of three spike results for each method (mean_%bioaccess/mean_{%spike recovery}) according to Koch et al. (2013).

In general, the main elements As, Pb, Fe, Ca, Cd, Cu, Zn, Mn, and S (Table I) and the compounds identified by XRD in soils match with the composition reported by Gutiérrez-Ruiz et al. (2003), Romero et al. (2008), Villalobos et al. (2010) and Martínez-Jardines (2018), for the same site. Ca and Pb arsenates (phamacolite and clinomimetite) were identified. They are secondary minerals that slowly form in soils of semi-arid regions. Anglesite (PbSO₄), arsenolite (As₂O₃), and clay with Fe, Ca, and Mg were also identified. CaCO₃ was detected indirectly in C3 through the reaction of carbonates with HCl. Nevertheless, other compounds reported in this terrain’s soil samples were not identified, possibly due to low concentrations or crystallinity. Most compounds found in analyzed soils in the study site included traces of arsenopyrite (Pokrovski et al. 2002), PbS, CaSO₄.2H₂O, goethite (FeOOH), magnetite (Fe₃O₄), chalcopyrite (CuFeS₂), and pyrite (FeS₂).

Furnace and converter dust and black arsenic must be the primary sources of As and Pb in the soil for two reasons: they are very rich in these elements, and calcine could contribute to Pb concentration (Table I). Overall, composites composition is similar to that of Cu slag (Nazer et al. 2016). The low geoavailability of Ca and Fe with high total concentration, possibly can be explained considering that these elements are in the slag. Moreover, variability of the total and geoavailable concentrations of minor elements (Table I and Fig. 2) shows a possible mixture of slag with converter flue dust, smelting furnace flue dust, sulfates, and carbonates. The diverse geoavailability of all elements except Fe (that was always low) can be attributed to the varied compositionof the residues and byproducts that polluted the soil as converter flue dust, smelting furnace flue dust, calcine, and black arsenic (Table I). The CaCO₃ in sample C3 explains the pH > 7. Pb, Cu, and Zn sulfates explain the pH < 7 of samples C1 and C2, as well as their low solubility in meteoric water (geoavailability) and high bioaccessibility in the gastric phase. These observations coincide with the data reported in other studies (Walraven et al. 2015, González-Grijalva et al. 2019). Despite this, not all of these metals must be sulfates. A portion of low concentrations may come from the acid lixiviation of slag in the gastric phase. Ca content in carbonate and plagioclase minerals could explain low geoavailability in meteoric water and high bioaccessibility under acidic conditions.

Arsenic geoavailability varies from 480 to 2855 mg/kg. Sample C3, with the lowest total As concentration and high geoavailability, emphasizes the great importance of the compound’s behavior. The soils contain As (III) in addition to As (V) compounds, which have been identified in this and other studies (Martínez-Jardines 2018). They can coexist because of dry weather conditions on site, and because reduction from arsenate to arsenite is slow (ATSDR 2007). The geoavailable As fraction must be mainly related with arsenolite, since the As (V) compounds have limited solubility in meteoric water. Although a number of arsenates could be adsorbed in amorphous Fe compounds (Goldber 2002, Donahoe et al. 2005, Hernández et al. 2016), it is not likely in this case. Fe in these soils seems to mainly be in the slag, because the geoavailable and bioaccessible fractions are quite low (Tables I, II). Furthermore, supposing the existence of amorphous oxides, Fe should be released during gastric extraction by chelation with organic ligands (Sidhu et al. 1981). Nevertheless, gastric Fe bioaccessibility was very low.

Pb geoavailability at pH = 5.5 was negligible but not its bioaccessibility, suggesting that Pb is mainly identified in low water-soluble sulfates, arsenates, carbonates and slag, but not under acidic conditions. It could also be identified in sulfides (González-Grijalva et al. 2019). All the compounds reported above can be partially solubilized under the acid condition of the gastric phase, even the slag (Aoki et al. 1984). Pb adsorption in birnessite or other Mn oxides that cover clays should be considered, since Mn compounds present hydroxyl groups favoring adsorption (Yin et al. 2011) at pH = 4.5-5.5 (Lenoble et al. 2002), although it is unlikely under gastric conditions. González-Grijalva et al. (2019) reported a Pb range from 40.8 to 50.8 %, lower than those obtained in this study. Considering that particle homogenization, temperature, and extraction times were the same, a possible factor related to data variability and soil composition could be the agitation system. The end-over-end shaker used in the NOM-147 method causes particle collision, reduction in particle size, and can increase soil reactivity (Quiroz-Vivanco 2018). The GSRSB uses a gas system that avoids biases in the expected results related to mechanical agitation. However, evidence of agitation’s influence is limited due to the insufficient number of analyzed samples.

Indeed, compound characteristics seem to be the most relevant factor explaining the variability. Even As lixiviated from slag can form soluble compounds under gastric conditions, depending on each reaction’s kinetics. Arsenic can exist as H₃AsO₃ and H₃AsO_4, or possibly as arsenates adsorbed in clays due to slow reduction. Clays do not dissolve in HCl (Simon and Anderson 1990), but can be partially altered, producing amorphous solids with a positive charge (Smedley and Kinniburg 2002). Acidic conditions increase the sorption capacity of clays, forming complexes with anions as arsenates at low pH (Simon and Anderson 1990, Magalhães 2002, Zhang and Selim 2005, Elsheikh et al. 2018). The soluble As (III) from arsenolite cannot be absorbed because the H₃AsO₃ is neutral and only acquires a negative charge at very high pH (Wang and Mulligan 2006).

Pb could be soluble or coordinated with organic ligands in the gastric solution. The NOM-147 method uses glycine, which forms Pb (NHCH₂COOH)₂ (Zhang et al. 2011), a soluble compound that mobilizes Pb from bones (Alcaraz-Contreras et al. 2011). The PBET method uses citrates, acetates, and other organic chelates. All of these can form soluble compounds with metals, releasing arsenates (Ruby et al. 1993). For example, citrates are useful to recover Pb from batteries (Villa-Vargas 2017, Villa et al. 2018), and acetates leach Pb from calcines (García-Villa 2016).

The observed differences between gastric and intestinal bioaccessibility have been related to particle size, mineral solubility, sorption complexes, soil characteristics, and new compounds (Ruby et al. 1999, Walraven et al. 2015). In this case, in soils with acid pH and low Ca content, a higher As bioaccessibility in the intestinal phase than in the gastric phase was observed. One possible explanation is the adsorption of arsenates on clays under gastric conditions, being released in the intestinal phase when the pH increases to neutrality with NaHCO₃ and the clays lose their charge. Although Singh et al. (2011) reported that some abiotic or biotic oxidoreduction reactions could happen with the added reagents changing the solubility of As, García-Rodríguez (2017) reported that adding only Na₂HCO₃ also recovered lower As in the gastric phase than in the intestinal, reinforcing our hypothesis.

Pb in the intestinal phase was lower than in the gastric phase, as has been reported by Yan et al. (2016). This is possibly because the Pb released from organic complexes precipitate at the intestinal pH (Ruby et al. 1993, 1999, Li et al. 2014). However, Pb (II) can be bio-transformed, modifying its solubility and toxicity (Calatayud and Llopis 2015, Cangelosi et al. 2017).

An increase of Pb bioaccessibility in the intestinal phase can enhance its toxicity. According to Kan et al. (2017), the Pb carbonates’ transformation to soluble organic Pb-complexes at neutral pH explain the increase of Pb bioaccessibility.

The bioaccessibility data for NIST 2710 in the gastric phase with the GSRSB- PBET method was more precise than with NOM-147, but this tendency is not clearly observed in the composite samples as mentioned above (Table II). Nevertheless, all the As and Pb values fall into the in vivo range reported by Koch et al. (2013) (Fig. 3), indicating that the reagents -and possibly the novel reactor system that reduces soil particle erosion and provides easier pH control- improve results.

In addition to those reported by several laboratories (Koch et al. 2013), all the data obtained in this study in the intestinal phase were lower than the in vivo range. It is worth mentioning that, according to Dabin et al. (2012), the in vitro methods with results similar to “in vivo” have longer agitation times or use more violent shakers. Therefore, it is possible that the intestinal phase results obtained with the GSRSB-PBET method can be improved by increasing agitation time.

The PBET method using the novel reactor system (GSRSB) provides easier pH control, and gastric phase data are more precise (% RSD) than the standard Mexican NOM-147 method.

The NIST 2710 results using the GSRSB-PBET method with the novel reactor show no statistical difference with respect to the in vivo measurements in the gastric phase.

The NOM-147 method measuring As and Pb bioaccessibility in the gastric phase does not simulate in vivo values and seems to overestimate the associated risk. Therefore, GSRSB-PBET method could be considered a precise alternative to measure As and Pb bioaccessibility in polluted soils modifying NOM-147 method and including the bioaccessibility measurement of other elements, mainly As.

The bioaccessibility of Pb measured during the intestinal phase (PBET) is always lower than in the gastric phase using both methods. Nevertheless, As behavior is the opposite except in C3, the only basic sample. It is recommended to include both phases and not only the gastric. It can also be possible to use the highest bioaccessibility value from both phases or perform in vivo bioavailability measurements.

Nevertheless, variability in soil results within the same metallurgical complex poses a challenge: to identify a general method with which to measure bioaccessibility in all types of contaminated soils, simulating in vivo data. Thus, it is essential to perform a complementary study to measure the bioaccessibility of a higher number of previously characterized soils. It must include at least the two studied methods and the two agitation systems with different agitation times and in vivo bioavailability determination.

We thank Daniel Amaro-Ramírez and Leticia Pérez-Manzanera for general support. The authors acknowledge Gerardo Martínez-Jardines and Teresa Pi-Puig (Instituto de Geología, UNAM) for technical support. T. García-Rodríguez is thankful to CONACYT Mexico for the fellowship received to perform Ph.D. studies.