CHAPTER 5: DISCUSSION
5.1 Reporter gene assays
Since the exact make-up of the leachates is unknown, only assumptions can be made regarding the content based on known endocrine disruptors found in and on plastic. Therefore, this discussion will focus on the expected effects of BPA, phthalates, BFRs, nonylphenols, POPs and, PAHs.
5.1.1 Effects on the AR
5.1.1.1 Anti-androgenic effects
There were no antagonistic effects evident in the AR inhibition assays. However, when BPA was investigated for its ability to interact with the androgen receptor it was found that BPA commonly acts as an AR antagonist. These results were determined using the African monkey kidney cell line CV-1 (Xu et al., 2005) and the MCF-7 cell line (Park et al., 2020). Furthermore, nonylphenols (Xu et al., 2005), phthalates (Christen et al., 2012), and BFRs (Hamers et al., 2006) were all found to exhibit antagonistic effects on the AR using CV-1, MCF-7, and AR-CALUX cells, respectively.
Although most of these studies were conducted on cells other than the MDA-kb2 cells, the lack of inhibition is not consistent with what is expected according to the literature. Therefore, the possible agonistic effects need to be considered to give insight into the overall effect of the leachates on the receptor.
5.1.1.2 Androgenic effects
There was no activation of the AR recorded in this study. As mentioned previously BPA commonly acts as an AR antagonist, and not an agonist (Xu et al., 2005; Park et al., 2020). Nonylphenol was also found to only inhibit the AR and not activate it (Xu et al., 2005). Also, a study done on 22 various phthalates, using the Chinese hamster ovary CHO-K1 cells, found no correlation between any of them and androgenic activity (Takeuchi et al., 2005). These studies were not conducted using the MDA-kb2 cell line, however. Therefore, the lack of androgenic response was to be expected regarding the assumed additives in the plastic.
It can be assumed that no androgenic adsorbents from the environment leached from the PVC.
This is because since the virgin plastics have not been exposed to the environment only the recycled plastics would have shown activation or inhibition. There is, however, no difference between the responses from the various PVC. A possible reason for the lack of expected results can be that the cell sensitivity between the assays differs. There are however no studies using the MDA-kb2 cell line for these purposes. Another likely reason that no effects were seen was
due to the mixture effects of the leachates. Mixture effects will be discussed further in this discussion.
5.1.2 Effects on the ER
5.1.2.1 Anti-oestrogenic effects
There was no measurable inhibition of the ER which was evident by the lack of dose responses.
However, unlike with the responses of the AR, the leachates exposed to the ER produced less light than the positive control in a few instances. For example, PVC A leached for 24 h at 4°C (Figure 4-16). This could be an indication of competition between the compounds that were leached. However, it is no easy feat to understand fully the oestrogenic activity of many of these compounds. Investigating only two of the additives made this evident, as bisphenol A has low estrogenic effects in vitro (Sharpe, 2010). However, in stimulating some cellular responses it can be found to have a similar potency to oestradiol (Rubin, 2011). Furthermore, when nonylphenols were tested for oestrogenic effects, it was determined that they were able to mimic oestrogen in only certain tissues (Watanabe et al., 2010). This increases the complication of determining the endocrine disruptive effects of these leachates. Therefore, it is important to consider the ingredients of mixtures that could occur, causing competition.
5.1.2.2 Oestrogenic effects
This study was unable to determine the ER activating effects of the leachates effectively due to too high a background of oestrogen activation existing within the cells. The activation measured after exposure to E2 alone was too high and not dose-responsive. There are a number of mechanisms of action leading to the expression of the target gene, particularly regarding the ER.
In an attempt to reduce the background activation of this assay, general laboratory protocol was set up to prevent any xenoestrogens from coming in contact with the cells. This included the use of higher-grade plastic that is less likely to leach; the use of HPLC water which is supposed to be without xenoestrogens; limited the use of EtOH for sterilization purposes because EtOH is known for ER interference (Singletary et al., 2001), and prevented ultraviolet exposure of any of the plastics that would come in contact with the cells. UV radiation is known to weather plastics increasing their potential to leach (Lee et al., 2012). These precautions had a notable effect on the response of the assay, however not enough of an effect to consider the assay working. As a last resort, ICI was added to the media when seeding the assay plates to allow for binding competition when the agonistic effect was evaluated. This was done to mimic what is usually done when setting up the AR inhibition assay that receives a background dose of testosterone. The addition of ICI was tested at various concentrations and given on days 1 and 2 of the assay respectively. With these results, the protocol used for this study was developed (See Materials
and methods 3.4.4). However, there was still too much background activation to make use of the assay for conclusive results.
The inability to complete the ER activation assay is regrettable, as many of the compounds that are assumed to leach are known ER agonists. Chemical analysis of the EDCs leached from marine microplastics found that ER agonists were present in the highest concentrations. Among all these BPA was found on 75% of the plastic, followed by bisphenol S (68%), octylphenol (63%), and nonylphenol (49%) (Chen et al., 2019). When leachates from water bottles incubated at varying temperatures were tested on the T47D-KBluc cells oestrogenic activity was measured.
Upon chemical analysis, BPA was detected in the samples possibly contributing to the oestrogenic activity (Aneck-Hahn et al., 2018). Drinking water sampled from Pretoria and Cape Town (South Africa) was also tested on the T47D-KBluc cells. Targeted chemical analysis found BPA, DBP, DEHP, and E2 in the samples that elicited an oestrogenic response (Van Zijl et al., 2017).
Another class of EDCs are phthalates, which are also known to mimic oestrogens. There are many kinds of phthalates and several in vitro studies have found that phthalates such as DBP can bind directly to the ER mediating gene expression (Takeuchi et al., 2005). A study using MCF- 7 cells, which are also human breast cancer cells, found that benzyl butyl phthalate, DBP and DEHP have oestrogen mimicking potential even at low concentrations (Chen & Chien, 2013). As previously mentioned, DBP and DEHP were found in drinking water that elicited an oestrogenic response using the T47D-KBluc cells (Van Zijl et al., 2017).
The oestrogen activity of BPA and phthalates alone suggests that the leachates may have adverse effects as the activity of the known endocrine disruptors, POPs, PAHs, and BFRs have not even been considered in this discussion.
5.1.3 Binding competition of mixtures
Most of the methods of risk assessment of EDCs focus on the effects of the individual chemicals.
However, chemical mixtures may have unexpected effects (Fent et al., 2006). An almost infinite number of combinations of contaminants are possible in environmental samples. Therefore, even the studies that have aimed at investigating the mixture effects of two or more chemicals are unable to determine accurately said effects in real-life scenarios (Carpenter et al., 2002). There are two theoretical models of chemical mixtures, interactive and non-interactive. Understanding how chemical mixtures react can lend to the understanding of why the effects seen in these results were not expected. It is often assumed that as long as the level of individual chemicals falls below a predetermined threshold, that the mixture will be safe. However, the interaction of the components within the mixture may have a far greater or smaller effect than what is expected
(Heys et al., 2016). An in vivo study was done to investigate the endocrine disrupting effects of a mixture of high molecular mass DEHP, low molecular mass DBP, and BPA. Each compound’s effect on relaxin family peptide receptor (RXFP2) signaling, which is responsible for the descent of testis in rats, was tested both individually and in a mixture. Whilst the individual compounds had an agonistic effect, increasing the response, surprisingly the mixture had the opposite effect (Suteau et al., 2020). Antagonistic mixture effects take place at the target site as the compounds compete for binding. This may be why there was no response from the AR.
A diverse group of environmental contaminants can act as AR antagonists by competitive or non- competitive binding to the AR (Hotchkiss et al., 2008). The mechanism of androgen action relies on the high-affinity binding of a ligand to the AR that causes a conformational change. This is required to prevent proteolytic degradation, and for AR dimerization and transcriptional activation.
In contrast, unliganded ARs are easily and rapidly degraded. AR antagonists, however, bind to the AR with low to moderate affinity. This may lead to the induction of a conformational change that is not able to protect the AR from degradation, or that is not compatible with ARE binding (Kelce & Wilson, 1997). It can be assumed that the compounds within may have a number of different binding affinities. Based on the known mechanisms, the lack of AR response altogether is, therefore, likely due to competition at the receptors.
The competition, as well as the values greater than 100% seen in the ER inhibition assay, can be due to any one or a combination of concentration addition, independent action, synergism, or potentiation. Synergism and potentiation are more complex to determine, as the effect of the mixture is far greater than that of the sum of the effects of the individual compounds. This occurs by either the induction or inhibition of detoxifying enzymes (Heys et al., 2016). The interaction involving these enzymes can occur not only between the compounds in the mixture but between the compounds and endogenous substrates as well (Hernandez et al., 2013).
Unlike the AR, there are several ER signaling pathways other than the classic E2-ER binding to the ERE. Nuclear E2-ER complexes can also be tethered through protein-protein interactions to a transcription factor complex. Growth factors can activate protein-kinase cascades, which lead to the activation of nuclear ERs. Finally, there are non-genomic actions whereby phosphorylation activates transcription factors (Bjornstrom & Sjoberg, 2005). Theoretically, the mixture of chemicals within the leachates could interfere with any of the above-mentioned mechanisms.
There is even the possibility that all the mechanisms are interfered with simultaneously (Carpenter et al., 2002).
There are a vast number of possible compound mixtures and various mechanisms of action.
Without chemical analysis and further testing of specific mixtures it is not possible to properly understand the ER inhibition results. Likewise, it is even less likely to understand the overall
adverse effects of the leachates without further testing. These tests could include the use of various cell lines with different sensitivities, and receptors. However, further tests should also include whole organism models, as there is currently a lack of sufficient knowledge to close the gap between in vitro and in vivo methods. Research such as this, however, aim to close that gap.