Fate of Bisphenol A in Terrestrial and Aquatic Environments

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Fate of Bisphenol A in terrestrial and aquatic environments Jeongdae Im, and Frank E Löffler Environ. Sci. Technol., Just Accepted Manuscript • DOI: 10.1021/acs.est.6b00877 • Publication Date (Web): 12 Jul 2016 Downloaded from http://pubs.acs.org on July 13, 2016

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Fate of Bisphenol A in terrestrial and aquatic environments

2 Jeongdae Im1 and Frank E. Löffler2,3,4,5*

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Department of Microbiology, University of Massachusetts,

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Amherst, MA 01002, USA 2

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Center for Environmental Biotechnology, University of Tennessee,

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Knoxville, TN 37996 3

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Department of Microbiology, University of Tennessee,

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Department of Civil and Environmental Engineering, University of Tennessee,

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University of Tennessee and Oak Ridge National Laboratory (UT-ORNL) Joint Institute for

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Biological Sciences (JIBS) and Biosciences Division, Oak Ridge National Laboratory,

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Oak Ridge, TN 37831

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* Corresponding author

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University of Tennessee

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Department of Microbiology

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M409 Walters Life Science Bldg.

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Knoxville, TN 37996

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Phone: +1-865-974-4933

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Fax:

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E-mail: [email protected]

+1-865-974-4007


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Abstract

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Bisphenol A (2,2-bis[4-hydroxyphenyl]propane, BPA), the monomer used to produce

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polycarbonate plastic and epoxy resins, is weakly estrogenic and therefore of environmental and

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human health interest. Due to the high production volumes and disposal of products made from

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BPA, polycarbonate plastic and epoxy resins, BPA has entered terrestrial and aquatic

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environments. In the presence of oxygen, diverse taxa of bacteria, fungi, algae and even higher

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plants metabolize BPA, but anaerobic microbial degradation has not been documented. Recent

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reports demonstrated that abiotic processes mediate BPA transformation and mineralization in

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the absence of oxygen, indicating that BPA is susceptible to degradation under anoxic

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conditions. This review summarizes biological and non-biological processes that lead to BPA

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transformation and degradation, and identifies research needs to advance predictive

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understanding of the longevity of BPA and its transformation products in environmental systems.

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Occurrence of BPA in the environment

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Bisphenol A (2,2-bis[4-hydroxyphenyl]propane, BPA) is used in the plastics industry as a

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monomer for producing polymeric materials, primarily epoxy resins, and polycarbonate plastic,

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and is also used as a raw material for other products such as flame retardants.1–3 The global

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demand of BPA exceeded 6.5 million tons in 2012 and is predicted to grow at an annual rate of

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4.6% from 2013 to 2019.4 BPA products have permeated the daily lives of people in many ways.

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For example, epoxy resins are used as dental sealants as well as internal protective coatings for

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food cans, bottle tops and water pipes, while polycarbonate is used in a wide variety of common

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products, including in digital media (e.g. CDs, DVDs), electronic equipment, automobiles, and

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medical devices. Also, in thermal papers, free or non-polymerized BPA is used, where one side


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has a powdery layer of a thermally reactive coating containing BPA (up to 2.3% by weight),5 and

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BPA contamination of recycled paper can occur.5,6 Due to the frequent and widespread use and

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disposal of these products, BPA is released into the environment, mainly through processing of

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BPA in manufacture, inefficient removal during wastewater treatment,7–10 landfill leachates,11–14

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and leaching from discarded BPA-based materials (e.g., hydrolysis of polycarbonate, recycled

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paper)5,15–17 (Figure 1). Consequently, BPA has been detected in various environmental systems

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(Table 1), and, according to a recent survey, BPA and three other bisphenols have been detected

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in human urine samples at high (up to 99%) frequency.18 Because of health concerns from

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exposure to BPA,19–23 further evaluation of the environmental fate of BPA is warranted.

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Currently, no data exist to support that BPA is produced naturally; however, there is evidence

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that bisphenol F (BPF), which is structurally very similar to BPA and may have comparable

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estrogenic potency24–26, is not a xenobiotic. For example, BPF and/or BPF derivatives were

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detected in orchids27–29 and bamboo shoots.30 In addition, BPF concentration averaging 3.2 mg

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kg-1 were measured in mild mustards made from Sinapis alba seeds.30 This is a remarkable

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observation because mustard has been used as a condiment since Roman times31, and with an

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average daily consumption of 1-2 g per person in the U.S. and Europe, the annual average BPF

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intake via mustard was estimated to range between 1.2 and 2.4 mg per capita.30,32

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Physicochemical properties of BPA

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The reported aqueous phase solubility of BPA is 300 mg L-1 at 25 °C,33 with an octanol-water

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partitioning coefficient (log Kow) of 3.42,34 and a Henry’s constant of 1.0

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These parameters indicate that BPA is not volatile at ambient temperatures but has a high

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tendency for sorption to soil and sediment.35–39 In addition, bound residue formation (i.e.,


10-10 atm m3 L-1.1

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incorporation of BPA into solid matrices involving covalent bonds) has been demonstrated.35,38

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Therefore, desorption/release from soil and sediment may be a key factor controlling transport,

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transformation, degradation and fate of BPA in the environment. The presence of heavy metals

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and cationic surfactants can enhance desorption of BPA from soils,40 and complete desorption

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can be accomplished, at least under laboratory conditions.40,41 Knowledge about the rates of BPA

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desorption from a variety of matrices are largely unknown but obviously relevant because

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aqueous phase BPA is more likely to be subject to natural attenuation including physical,

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chemical, and biological processes that reduce contaminant toxicity.42,43

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Scope

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A PubMed search with the term “bisphenol A” (May 2016) identified more than 10,000

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publications and reports investigating possible human health effects, and comprehensive review

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articles discuss human exposure5,44,45 and toxicological aspects of BPA.3,19,20,46 Several national

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health risk assessment reports are also available.21–23 This review does not discuss routes of

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exposure to BPA and toxicological aspects. Rather, the aim is to assess the contributions of

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biological and non-biological processes to BPA transformation and degradation in environmental

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systems, and to summarize available process-specific information. Gaps in our understanding of

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BPA degradation are identified with the goal to focus future research efforts and establish a

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scientific basis for predicting the fate and longevity of BPA. Such information is needed to give

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regulatory agencies the means for establishing meaningful maximum concentration level (MCL)

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values and for informing the concerned public about possible exposure routes and realistic

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dangers associated with environmental BPA. Figure 1 illustrates major sources and sinks of


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environmental BPA, as well as recognized natural attenuation processes, which are the focus of

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this review.

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Aerobic bacterial metabolism

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Oxic conditions. In the presence of oxygen, BPA is susceptible to bacterial degradation, and

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numerous bacterial isolates belonging to the α-Proteobacteria, β-Proteobacteria, γ-Proteobacteria,

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Bacillus and Actinobacteria (Table 2) have been obtained and characterized. The first bacterial

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isolate demonstrated to degrade BPA was Sphingomonas sp. strain MV1 isolated from sludge

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collected from a wastewater treatment plant at a plastics manufacturing facility.47 Initially

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defined as a coherent group of Gram-negative, strictly aerobic bacteria,48 the genus

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Sphingomonas was later subdivided into the genera Sphingomonas, Sphingobium,

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Novosphingobium, and Sphingopyxis within the family Sphingomonadaceae.49 The BPA-

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degrading phenotype has been observed in all genera of this family,50–52 whose members

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commonly occur in soil,53–56 freshwater,48 seawater,57 and wastewater47 habitats. As of 2015, 22 genera within the classes a-, b-, g-Proteobacteria, Bacilli, and

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Actinobacteria have been reported to harbor members capable of metabolizing BPA under oxic

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conditions (Table 2). Apparently, diverse bacterial groups inhabiting various environments share

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the capability of metabolizing BPA in the presence of oxygen. Based on the identification of

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BPA degradation intermediates, several pathways have been proposed. The major metabolic

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routes proceed via oxidative skeletal rearrangement producing 4-isopropenylphenol (IPP), 4-

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hydroxybenzaldehyde (HBAL), 4-hydroxybenzoate (HBA), 4-hydroxyacetophenone (HAP), 4-

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hydroxycumyl alcohol (HCA) and hydroquinone (HQ), which represent the most frequently

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observed BPA metabolites (Figure 2).47,58–60 IPP, HBAL, HBA, HAP, HCA, and HQ are


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mineralized and assimilated into cell carbon.47,58 An alternative, presumably minor route,

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involves an initial hydroxylation step to produce 2,2-bis(4-hydroxyphenyl)-1-propanol, which is

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further oxidized to HBA and 4-hydroxyphenacyl alcohol.47,58–61 The application of high-

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resolution mass spectrometry techniques allowed the identification of other, presumably minor

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intermediates62–65, and more complicated pathways that may contribute to BPA degradation have

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been proposed.62–66

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Bacterial genes and enzymes involved in aerobic BPA metabolism. Aerobic BPA

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degradation pathways have been exhaustively studied, but relatively little is known about the

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catalysts (i.e., enzymes) and the genes (Table 3). The addition of metyrapone, a cytochrome

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P450 inhibitor, decreased BPA degradation activity of Sphingomonas sp. strain AO1.53

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Subsequently, a cytochrome P450 monooxygenase system, consisting of cytochrome P450 and a

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ferredoxin, was confirmed to be involved in BPA degradation using enzyme preparations of

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strain AO1.54 Genome walking identified bisdA and bisdB genes encoding ferredoxin and

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cytochrome P450, respectively, in strain AO1.67 Subsequently, Escherichia coli harboring a

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bisdAB-recombinant plasmid was demonstrated to hydroxylate BPA, indicating that the bisdAB

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genes were responsible for initiating BPA degradation in strain AO1.67 E. coli harboring only

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bisdB also hydroxylated BPA, suggesting that native E. coli proteins fulfilled the ferredoxin

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electron transfer function enabling cytochrome P450 activity in the E. coli mutant.67 Cytochrome

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P450 monooxygenase was proposed to mediate an ipso substitution mechanism, whereby BPA is

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initially hydroxylated at the ipso position to form a quinol intermediate.62–64 Following

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rearomatization, the C-C bond between the semiquinol and the isopropylphenol moieties is

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cleaved to produce hydroquinone and carbocationic isopropylphenol.63 More recently, BPA

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metabolism in Sphingobium sp. strain BiD32 was investigated using quantitative proteomics and


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metabolomics.68 BPA-M (1,2-bis(4-hydroxyphenyl)-2-propanol) was proposed as the first

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intermediate of BPA degradation.68 Among the proteins expressed in response to the presence of

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BPA, a novel p-hydroxybenzoate hydroxylase was implicated in the initial BPA transformation

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step and BPA-M formation.68 A putative cytochrome P450 encoded on the Sphingobium sp.

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strain BiD32 genome was not detected; however, a ferredoxin, which is part of the cytochrome

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P450 monooxygenase system, was measured. The ferredoxin was not differentially expressed in

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response to BPA,68 suggesting that strain AO1 and strain BiD32 may use different BPA

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degradation pathways. In summary, a number of bacterial species degrade BPA in the presence

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of oxygen, suggesting that aerobic processes contribute to BPA attenuation in the

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environment.1,66,69,70 Although some pathway information has emerged, additional efforts are

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warranted to elucidate the genetic basis (i.e., genes), and the mechanisms (i.e., enzymes) that

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govern bacterial BPA degradation in oxic environments.

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Bacterial co-metabolism

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The initial microbial transformation of BPA does not always lead to productive degradation, and

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the benefits to the organisms are not apparent (i.e., co-metabolism). For example,

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Mycobacterium sp. catalyzed O-methylation of BPA under oxic conditions, leading to the

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formation of monomethyl- and dimethyl ethers (BPA-MME and BPA-DME, respectively)

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(Figure 2).71 O-methylation diminished the affinity of BPA for the estrogen receptor,72 but

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increased toxicity.71–73 Microbial degradation of BPA-MME and BPA-DME has not been

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explored and the fate of these ether compounds is uncertain. The magnitude of O-methylation

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activity is unknown and further investigations are needed to determine the occurrence and

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environmental behavior of these methylated BPA transformation products.


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BPA disappearance was observed in Nitrosomonas europaea cultures expressing

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ammonia monooxygenase (AMO), whereas no BPA loss was observed in control incubations

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amended with allylthiourea, a potent inhibitor of AMO activity.74 Similar observations were

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made in activated sludge incubations, and the BPA concentrations decreased concomitant with

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ammonia oxidation, but ceased in the presence of allylthiourea.75 No reaction intermediates or

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degradation end products were identified. An independent study identified nitro- and dinitro-

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BPA in N. europaea cultures amended with BPA at pH 6.0, and cell-free incubation

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demonstrated the abiotic nitration of BPA in the presence of nitrite (Figure 2).76 Hence, it is

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currently unclear if AMO is directly involved in BPA transformation or if BPA reacts chemically

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with the nitrite generated in the AMO reaction. The fate of nitrated bisphenols is currently

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unclear but nitration apparently reduced estrogenic activity.76 Nitration was not observed at pH 7

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and 8 suggesting this process may be limited to acidic conditions.

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Methanotrophic bacteria possess soluble methane monooxygenase (sMMO) and

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particulate methane monooxygenase (pMMO), both enzyme systems with broad substrate

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specificities that fortuitously attacks an array of organic compounds.77–80 Methylosinus

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trichosporium strain OB3b possesses both monooxygenases and serves as a model organism for

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evaluating methanotrophic co-metabolism.81 Experiments with strain OB3b failed to demonstrate

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BPA transformation under conditions conducive for sMMO and pMMO activity (Im et al.,

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unpublished data). The current understanding of co-metabolic transformation potential of

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bisphenols is limited, and additional testing of a broader diversity of ammonia-oxidizing and

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methane-oxidizing microorganisms is warranted.

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Some pseudomonads (e.g., Pseudomonas sp. strain LBC1) produce extracellular laccases capable of catalyzing BPA transformation reactions (Table 3).82 The strain LBC1 laccase


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efficiently degraded BPA even without redox mediators, which facilitate electron transfer and

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can enhance laccase activity.82,83 These examples demonstrate that taxonomically diverse

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bacteria expressing distinct enzyme systems that share broad range substrate specificity can co-

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metabolize BPA under oxic conditions. Of note, BPA co-metabolism in the absence of oxygen

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has not been reported and evidence for anaerobic co-metabolic transformation and degradation of

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BPA is lacking.

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Fungal transformation and degradation

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Enzymes produced by ligninolytic fungi, including lignin peroxidase (LiP), manganese

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peroxidases (MnP), versatile peroxidases (VP), and laccases are nonspecific, and therefore have

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attracted attention for initiating degradation of many recalcitrant organic pollutants.83–88 Fungi

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and fungal enzymes demonstrated to attack BPA are summarized in Table 3. The white-rot

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basidiomycete Pleurotus ostreatus degraded BPA, and in vitro experiments demonstrated that

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the organism’s MnP was involved in phenol, IPP, 4-isopropylphenol, and possibly hexestrol

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formation (Figure 2).89 Similarly, MnPs produced by Phanerochaete chrysosporium and

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Trametes versicolor quickly and completely removed BPA from the culture medium in

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laboratory tests, and estrogenic activity declined with extended incubation times.90 These results

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suggested that potential BPA transformation products with estrogenic activity were also

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susceptible to MnP-mediated degradation.

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In general, fungal laccases are less effective for BPA transformation and degradation than

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MnP, but redox mediators can enhance their reactivity. For example, the presence of 1-

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hydroxybenzotriazol,90,91 2,2-azino-bis(3-ethylbenzothiazoline-6-sulphonic acid),92,93 or

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acetosyringone93, fungal laccases demonstrated effective BPA transformation. The purified


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laccase from Trametes villosa transformed BPA to the mono-phenolic compound IPP and a BPA

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dimer, suggesting that laccase activity may result in successive BPA polymerization.94 Similar

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laccase-mediated BPA oligomerization was observed in cultures of the white rot fungus

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Coriolopsis polyzona and BPA dimers, trimers, and tetramers were detected.92 The laccase of

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Pycnoporus coccineus oxidized nonylphenol, octylphenol, and ethynylestradiol, but not BPA,95

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indicating that not all fungal laccases initiate BPA transformation.

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Algal BPA transformation

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Microscopic, photosynthetic microalgae have been extensively investigated for bioremediation

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applications as microalgae can sorb, accumulate, and/or metabolize a variety of contaminants.96

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Under photoautotrophic conditions, Chlorella fusca and Anabaena variabilis removed 85% ± 7%

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and 23% ± 6% of BPA, respectively, in 120 hours from aqueous solution containing 40 µM BPA

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but no degradation products were detected (Hirooka et al., 2003).97 The microalgae

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Nannochloropsis sp. and their zooplanktonic predators Artemia sp. and Brachionus sp. were

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employed to investigate the possible accumulation of BPA along the food chain.98 These efforts

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revealed that almost all BPA remained in the liquid medium when the zooplankton was

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incubated alone. In contrast, 40% of the BPA was recovered from the zooplankton when

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incubated together with the microalgae, suggesting BPA transfer and accumulation across

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trophic levels.98 The major BPA transformation products formed by the green algae

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Pseudokirchneriella subcapitata, Scenedesmus acutus, Scenedesmus quadricauda, Coelastrum

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reticulatumwere, and Pavlova sp. were identified as BPA-glycosides, such as BPA-mono-O-b-D-

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glucopyranoside and BPA-mono-O-b-D-galactopyranoside (Figure 2).99,100 Glycosylation of BPA

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to non-estrogenic BPA-glycosides has also been observed in fungi and plants, presumably as


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intermediates of a detoxification process.99–106 Although BPA-glycosides do not have estrogenic

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activity, hydrolases (i.e., b-glucosidase) can release BPA in the mammalian intestine.99 To date,

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no BPA degradation intermediates have been reported as a consequence of algal metabolism, at

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least in axenic cultures. Recently, symbiotic (i.e., commensal) BPA degradation was reported

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and the microalgae Chlorella sorokiniana provided oxygen to a BPA-degrading bacterial mixed

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culture.107

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Phyto-transformation and degradation of BPA

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In phytoremediation, green plants transform, stabilize, and remove inorganic and organic

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contaminants from soil and water.108 Several plant species have been investigated for their ability

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to remove BPA.102–104,109–121 Three main BPA metabolic pathways were observed in plants and

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plant extracts: (i) conjugation with soluble carbohydrates (glycosylation), (ii) formation of bound

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residues, and (iii) formation of polar polymers.100,102,104,110,111 Among these pathways, BPA

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glycosylation is regarded as the main metabolic route of BPA in plants. For example, Nicotiana

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tabacum cell suspensions102 and Ipomoea aquatica (water spinach)104 absorbed and metabolized

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BPA to BPA-mono-O-b-D-glucopyranoside as a major product, and several other BPA-

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glycosides were detected in Eucalyptus perriniana cell suspension cultures (Figure 2).111 The

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fate of these BPA adducts is unclear and it is possible that BPA-glycoside hydrolysis releases

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BPA.99 Therefore, further investigations of BPA metabolites in edible plants are warranted. A

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few studies demonstrated BPA degradation associated with plant metabolism, and monophenols,

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such as IPP and HCA, were reported as BPA metabolites in addition to monohydroxyl BPA and

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BPA-glycosides (Figure 2).111,117 Oxidative enzymes, such as peroxidases and polyphenol


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oxidases, have been implicated in the ring cleavage reactions118–122 (Table 3) suggesting that

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BPA degradation in plants could potentially occur.

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BPA degradation under anoxic conditions

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Information about microbial BPA degradation in the absence of oxygen is scarce, and several

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studies have concluded that BPA is recalcitrant to anaerobic microbial (co-) metabolism.37,39,123–

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1).127,128 No microbial BPA degradation was observed in anoxic microcosms established with

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freshwater sediment,36,39,124,125,129 marine sediments,126 and soil,37 leading to the conclusion that

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BPA is recalcitrant and undergoes little or no biodegradation in the absence of oxygen.

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Reductive dehalogenation of halogenated BPA derivatives occurred in estuarine sediment

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microcosms, but no further degradation was observed under anoxic conditions.125,130–135

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Contrasting these observations, a BPA half-life of less than 1 day was reported in denitrifying

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columns established with aquifer material.136 BPA disappearance was also reported under

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sulfate-, nitrate-, and iron-reducing conditions established in microcosms derived from

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freshwater sediment.137 Experiments with Bacillus sp. strain GZB, a facultative anaerobe,

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suggested BPA removal under both oxic and anoxic conditions.138 Although these studies

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demonstrated BPA disappearance, BPA transformation/degradation products were not measured

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and quantified, and BPA degradation in the absence of oxygen has yet to be demonstrated. Ten

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strains of Lactococcus (four L. lactis subsp. lactis, five L. lactis subsp. lactis bv. diacetylactis,

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and one L. lactis subsp. cremoris) were examined for their ability to remove BPA from aqueous

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solution under anoxic conditions.139 While the aqueous phase BPA concentrations decreased by

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9-62% in live incubations, no degradation products were detected. An independent study

This is of interest because a substantial mass of BPA resides in anoxic sediments (Table


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reported BPA concentration decreases of 20% in E. coli cultures unable to degrade BPA.140

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Additional experiments revealed that BPA adsorption to the biomass, rather than

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transformation/degradation, explained BPA disappearance.139,140 These are relevant observations

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indicating that BPA disappearance is a poor indicator of degradation and experimental data

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documenting a decrease in BPA concentrations must be carefully interpreted. Taken together,

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unequivocal evidence demonstrating metabolic and co-metabolic microbial BPA degradation in

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the absence of oxygen is currently lacking.

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Abiotic BPA transformation and degradation

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A number of abiotic processes can lead to BPA transformation and degradation. While processes

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involving reactive oxygen species (ROS) are generally limited to oxic environments, reactive

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mineral phases can mediate BPA degradation under oxic and anoxic conditions.

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Photo-degradation. BPA is susceptible to light-induced transformation. In photolysis (photo-

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oxidation), photons initiate a chemical breakdown process, which represents an important

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transformation pathway for many organic pollutants in surface waters and surface soils.141,142

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BPA half-lifes in surface waters ranged from 66 hours to 160 days and the efficiency of

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photolytic degradation depended on pH, turbidity, water turbulence and other factors.143 BPA

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photo-degradation intermediates include phenol, 4-isopropylphenol and a semi-quinone

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derivative of BPA.143

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Photo-oxidation, also known as indirect photolysis, refers to the degradation of a

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compound by naturally occurring ROS including hydroxyl radicals (OH•), peroxide radicals

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(ROO•) and singlet oxygen (1O2) generated by light. For example, in the presence of nitrite or

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nitrate, solar radiation can induce the formation of hydroxyl radicals from water,144 and the


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subsequent ROS-mediated transformation of BPA has been demonstrated.145,146 In the presence

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of ROS, NaCl enhanced BPA degradation by producing reactive hypochlorite (OCl-).15

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Dissolved organic matter,145,147,148 ferric iron,145,149 lipids,150 and riboflavin151,152 have also been

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implicated in ROS formation and BPA transformation.

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Advanced oxidation. BPA concentrations in wastewater effluent range from low ppb to several

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ppm levels (Table 1). Many studies demonstrated the utility of advanced oxidation (AO)

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processes based on H2O2- (e.g., Fenton chemistry), UV light-, and ozone-enabled removal of

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BPA in aquatic compartments. The interested reader is referred to review articles describing AO

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processes in detail.10,153–157 AO processes can effectively remove BPA in engineered systems

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(e.g., wastewater treatment plants), and possibly expanded for enhanced treatment of

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contaminants in aquifers and sediments. A recent study demonstrated that Shewanella oneidensis

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mediated Fenton chemistry-based degradation of 1,4-dioxane when the culture was provided

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with a suitable electron donor (lactate) and ferric iron as an electron acceptor, and was exposed

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to alternating oxic-anoxic conditions.158 This observation suggests that microbially mediated

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Fenton chemistry can potentially contribute to contaminant degradation at circumneutral pH near

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oxic-anoxic interfaces.

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Zero-valent iron (ZVI). ZVI is reactive towards a variety of organic compounds.159 The ZVI-

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mediated production of hydroxyl radicals was implicated in BPA degradation, but no

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transformation products were reported.160 Similarly, degradation of aqueous phase BPA using

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nano-sized ZVI in the presence of hydrogen peroxide and persulfate oxidants has been

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demonstrated.161

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Reactive mineral phases. Manganese and iron can form reactive mineral phases that affect the

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fate and transport of organic contaminants via sorption, hydrolysis and/or oxidative


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transformation. In particular, manganese oxides are strong, naturally occurring oxidants, and

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play relevant roles in the biogeochemical cycling of carbon and other elements.162–164 Recent

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studies demonstrated that manganese dioxide (MnO2)165–168 mediates BPA transformation.

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As a strong oxidant, MnO2 serves as an electron acceptor in microbial respiration under

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anoxic conditions and also as a mineral phase reactive towards many organic contaminants.165–171

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Oxidative transformation of BPA by MnO2 has been demonstrated using synthetic MnO2,165–168

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and several transformation pathways were suggested based on the identification of reaction

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intermediates.165 A proposed initial electron transfer step to MnO2 results in the formation of

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BPA radicals, which subsequently undergo a variety of reactions (Figure 2).165 HCA was

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identified as a major transformation intermediate, and at least 64% (mol/mol) of the initial

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amount of BPA was recovered as HCA (Figure 2).168 HCA was also susceptible to MnO2-

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mediated degradation and mineralization (i.e., CO2 evolution) occurred, indicating that complete

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destruction of BPA can be achieved in the reaction with MnO2 in the absence of oxygen, at least

335

under laboratory conditions.168 Compared to BPA, HCA has an octanol-water partition

336

coefficient (Log Kow) of 0.76 (BPA 2.76) and an aqueous phase solubility of 2.65 g L-1 (BPA

337

0.31 g L-1).168 The different physicochemical properties suggest that HCA has increased mobility

338

in water-saturated systems compared to BPA, and is therefore more likely to encounter oxic

339

zones. HCA appears recalcitrant to microbial degradation under anoxic conditions, but is rapidly

340

utilized as a growth substrate in the presence of oxygen.168 Recently, the utility of MnO2 as an

341

additive to engineered stormwater infiltration systems to oxidize organic contaminants, including

342

BPA, was demonstrated.172 Goethite also has been reported to mediate degradation of certain organic pollutants,173,174

343 344

and oxidative transformation of BPA has been observed in aqueous suspensions of goethite (a-


15

Page 17 of 44


345

FeOOH).175 Transformation reactions similar to those observed in MnO2-mediated BPA

346

degradation were proposed,175 but HCA was not detected, suggesting that the goethite-mediated

347

BPA degradation pathway differs.

348

Goethite is a common and thermodynamically stable iron oxide in soils,176 and

349

manganese oxides occur widely in freshwater and marine sediments. For example, 25 to 185

350

µmol mL-1 of MnO2 have been detected in marine sediments,177 and even greater amounts of 300

351

- 4,000 µmol mL-1 of MnO2 have been observed in freshwater sediments.178,179 The amounts of

352

MnO2 observed in natural soils and sediments are in the range of nominal MnO2 concentrations

353

used in laboratory experiments that demonstrated MnO2-mediated BPA mineralization.168 Of

354

course, many factors including pH, dissolved organic matter, metal ion concentrations, microbial

355

activity, etc., influence mineral phase formation, reactivity, and passivation165,180–182 and the

356

contributions of reactive mineral phases to BPA degradation in environmental systems have yet

357

to be evaluated. A relevant aspect that is poorly understood is the interplay between biotic and

358

abiotic reactions. Ferrous iron- and manganic ion-oxidizing microbes play key roles in

359

generating the reactive mineral phases, and more detailed studies of biologically mediated

360

abiotic degradation (BMAD) processes (e.g., microbial Mn2+ oxidation and mineral phase

361

formation followed by MnO2-mediated BPA degradation) are needed. It is conceivable that such

362

BMAD processes have major contributions for in situ BPA attenuation and for controlling the

363

fate and longevity of BPA in sediments and aquifers.

364 365

Bound residue formation

366

Many hydrophobic organic compounds and their transformation products form non-extractable

367

bound residues in soils,183 and this process has been considered a major attenuation pathway.184


16


Page 18 of 44

368

Bound residue formation of BPA into the soil matrix has been demonstrated under oxic

369

conditions, and the portion of the bound residue increased significantly as a function of time,35,38

370

and microbial activity has been implicated in bound residue formation.38,185,186 On the other

371

hand, a substantial amount of tetrabromobisphenol A (TBBPA) incorporated into the soil matrix

372

under anoxic conditions, but was released upon exposure to oxygen.38 The release of

373

contaminants covalently incorporated into the biomass has been reported in response to changing

374

physicochemical conditions and microbial activity.186,187 For example, a change in soil redox

375

conditions can impact the structure of natural organic matter (NOM), which affects the binding

376

and sorption of organic pollutants to NOM.38,188

377

Bound residue formation decreases the risk of exposure, but at the same time decreases

378

contaminant availability for biotic and abiotic degradation. There is an ongoing debate regarding

379

whether bound residues of organic compounds remain (bio)available in the long term.187,189

380

Further studies on the stability of bound residues of BPA in soils or sediments under dynamic

381

environmental conditions are warranted.

382 383

Knowledge gaps

384

BPA and BPA derivatives are, and will remain, part of modern societies for decades to come,

385

and so will the associated concerns, unless research provides clear understanding about the fate

386

and the longevity of bisphenols in environmental systems. As outlined above, a variety of biotic

387

and abiotic processes that contribute to BPA transformation, degradation, and incorporation into

388

solid matrices (e.g., soil, sediment) have been identified and documented in the laboratory. Still,

389

many open questions about the ultimate fate of BPA and its metabolites in environmental

390

systems remain, and more information about degradation pathways and mechanisms, cornerstone


17

Page 19 of 44


391

organisms, enzymes and genes, as well as favorable geochemical conditions that sustain

392

acceptable degradation rates is needed. The biogeochemical settings conducive for BPA

393

degradation and detoxification must be better delineated, and tools to quantitatively assess BPA

394

transformation/degradation processes are desirable. Under anoxic conditions, mineral phase-

395

mediated reactivity is relevant but the turnover and formation of the reactive minerals (e.g.,

396

MnO2) are not fully characterized, making predictive understanding of the contributions of

397

abiotic degradation tenuous. Most knowledge about the fate of BPA has accrued from

398

laboratory-based studies but comprehensive efforts to explore the relative contributions of

399

different attenuation mechanisms under in situ conditions are lacking.

400 401

Perspective on the fate of BPA in terrestrial and aquatic environmental systems

402

Identification of the major BPA sources is a key first step for addressing the concerns associated

403

with BPA. Among the sources of environmental BPA shown in Figure 1, permitted wastewater

404

effluent discharge is considered the main entry way for BPA into terrestrial and aquatic

405

environments.190,191 As outlined in this review, a substantial body of knowledge has accrued

406

about biotic and abiotic BPA transformation/degradation processes and pathways in laboratory

407

studies. Table 4 summarizes the reported processes and the associated rates leading to BPA

408

removal. Interestingly, the degradation/transformation rates observed in laboratory studies using

409

bacteria, fungi, algae and plants are in a similar range, and in situ studies are needed to decipher

410

the relative contributions of the different processes for BPA decomposition under

411

environmentally relevant conditions. Of note, these processes require oxygen, and under anoxic

412

conditions, BPA transformation/degradation appears to be limited to reactions with reactive

413

mineral phases. Laboratory studies reveal that MnO2-mediated BPA degradation occurs at


18


Page 20 of 44

414

several orders of magnitude higher rates than those reported for other processes (Table 4).

415

Considering that a major concern is BPA associated with (anoxic) sediment, Mn-cycling near

416

oxic-anoxic interfaces may be a major process controlling the fate of BPA in sediments. Based

417

on the current understanding of processes leading to BPA removal, engineering solutions at the

418

identified key points of entry could be envisaged to eliminate further release of BPA into the

419

environment. The information obtained from laboratory studies indicates that BPA is susceptible

420

to natural attenuation processes under oxic and anoxic conditions, and an important next step

421

will be to assess the relevance of the different potential attenuation processes on environmental

422

BPA concentrations. Understanding of BPA transformation and degradation has advanced to a

423

level that justifies integrated in situ studies that consider all known BPA attenuation

424

mechanisms. Comprehensive in situ efforts can evaluate the environmental fate of BPA and

425

reveal realistic exposure routes, so that potential dangers to human and environmental health can

426

be clearly delineated and predicted.

427 428

Acknowledgements

429

This work was supported by the Polycarbonate/BPA Global Group of the American Chemistry

430

Council (ACC), Washington, DC, USA.

431 432

References

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1000 1001 1002 1003 1004 1005 1006 1007 1008 1009 1010 1011 1012 1013 1014 1015 1016 1017 1018 1019 1020 1021 1022 1023 1024 1025 1026 1027 1028 1029 1030 1031 1032 1033 1034 1035 1036 1037 1038 1039 1040 1041 1042

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Page 36 of 44

1061 1062 1063

Table 1. Occurrence of BPA in the environment exemplified by representative published information. Extensive compilations detailing BPA occurrence in environmental systems are publicly available (e.g., refs 10, 192, 194–197 and NORMAN-EMPODATa).

1064

Source

Concentration (ng L-1)

References

Surface water

0 – 272

127

9 – 776

128

0.5 – 410

8

140 Seawater

b

192

81b (North America), 10b (Europe)

190

0 – 249

128

b

Sewage effluent

Sewage sludge

Soil/sediment

c

Landfill leachate

b

0 (North America), 1.6 (Europe)

190

0 – 2.6

9

31 – 223

7

18 – 702

8

0 – 32.4

9

0 – 17,300 (0 – 1,500)b

10

0 – 12,500

7

70 – 770

127

4 – 1,363

8

0 – 15

127

66 – 343

128

0.01 – 0.19

8

0.6/3.5b,d (North America), 16/8.5b,d (Europe)

190

0.5 – 325

196

1.3 - 17,200 6

4.2 × 10 – 25 × 10 17,000

11 6

a

13

10 – 107,000

1065 1066 1067 1068 1069

12 14

a

NORMAN-EMPODAT Database (http://www.normandata.eu/empodat/).197 NORMAN is a network of reference laboratories, research centers and related organizations for monitoring of emerging substances. EMPODAT is a database of geo-referenced monitoring and bio-monitoring data on emerging substances in the following matrices: water, sediment, biota, soil, sewage sludge and air. bmedian value; cunit: ng g-1 d dw ; values obtained from freshwater and marine sediments, respectively;


35

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1070


Table 2. Bacterial isolates capable of degrading BPA under oxic conditions and their origins. Bacterial isolates

Origin

a-Proteobacteria

Ensifer adhaerens strain J3 Sphingomonas bisphenolicum strains AO1, Sphingomonas sp. strains MV1, AO1, SO11, BP7, and TTNP3 Sphingobium yanoikuyae strain BP-11R and Sphingobium sp. strains BiD32 and YL23 Novosphingobium sp. strains FID3, TYA-1 Sphingopyxis sp. strain BiD10 Ochrobactrum sp. strain T Alcaligenes sp. strain OIT7 Bordetella sp. strain OS17 Achromobacter xylosoxidans strain B-16 Nitrosomonas europaea Pandoraea sp. strain HYO6 Cupriavidus basilensis strain JF1 Variovorax sp. strain HCA

FW SS, WTP, SW

198 42, 47, 53, 56, 57, 199

FW, WTP SS WTP EW SS SS CL SS SS WPT SS

50, 52, 200 51, 201 52 133 56 56 202 74 56 65

Serratia rubidae strain LCA3 and Serratia sp. strain HI10 Enterobacter gergoviae strain BYK-7, Enterobacter sp. HI9, HA18, BPR1, and BPW5 Klebsiella pneumoniae strain BYK-9, Klebsiella sp. NE2, SU3 and SU5 Aeromonas hydrophila Pseudomonas monteilii strain N-502, Pseudomonas paucimobilis strain FJ-4, Pseudomonas knackmussii strain B13, Pseudomonas putida strain KA5, Pseudomonas aeruginosa strains LCS1 and 2, Pseudomonas sp. strains LBC1, KA4, SU1, SU4, SU19, KU1, and KU2 Bacillus cereus strain BPW4, Bacillus amyloliquefaciens strain Bak15a, Bacillus sp. strains BP-2CK, BP-21DK, GZB, NO13, NO15, YA27, and KU3 Streptomyces sp. Mycobacterium vanbaalenii strain PYR-1 Arthrobacter sp. strain YC-RL1

SS, FW SS, WTP SS, WTP SS SS, WTP, SW, FW, UI

56, 203 56, 116, 204 56, 204 205 56, 59, 82, 123, 203, 206, 207, 208, 209

SS, WTP, SW, FF FW FW SS

56, 107, 138, 209–211

b-Proteobacteria

g-Proteobacteria

Bacillus Actinobacteria

1071 1072

a

Classes

References

b

212 71 213

a

SS: soil/sediment; WTP: wastewater treatment plant; SW: seawater; FW: freshwater; EW: electric waste recycling site; CL: compost leachate; UI: unidentified; FF: fermented food; bIm et al., unpublished.


36


1073 1074

Page 38 of 44

Table 3. Organisms harboring enzyme systems that catalyze BPA transformation reactions. Kingdom

Species

Enzyme

References

Bacteria

Sphingomonas sp.

Cytochrome P450

53, 54, 63

Sphingobium sp.

p-hydroxybenzoate hydroxylase

68

Pseudomonas sp.

Laccase

82

Nitrosomonas europaea

Ammonia monooxygenase

74

Pleurotus ostreatus, Phanerochaete chrysosporium, Trametes versicolor, Irpex lacteus

Manganese peroxidase

89, 90, 214, 215

Coprinus cinereus, Bjerkandera adusta

Peroxidase

216, 217

Trametes vasocolor, Trametes villosa, Phanerochaete chrysosporium, Irpex lacteus, Coriolopsis polyzona, Paraconiothyrium variabile, Grifola frondosa, Phlebia tremellosa, Aspergillus niger, Phlebia tremellosa

Laccase

94–97, 214, 215, 218– 223

Cunninghamella elegans, Phanerochaete chrysosporium

Cytochrome P450

224, 225

Unspecified fungus

Polyphenol oxidase

226

Horseradish, Soy bean

Peroxidase

118, 119, 121, 122

Potato

Polyphenol oxidases

120

Fungi

Plantae


37

Page 39 of 44

1075


Table 4. BPA degradation/transformation pathways and reported kinetics constants. Degradation/transformation mechanism Bacteriaa Achromobacter xylosoxidans Alcaligenes sp. Bacillus cereus Bacillus pumilus Bacillus sp. Bordetella sp. Cupriavidus basilensis Ensifer adhaerens Enterobacter gergoviae Enterobacter sp. Klebsiella pneumonia Klebsiella sp. Novosphingobium sp. Pandoraea sp. Pseudomonas aeruginosa Pseudomonas knackmussii Pseudomonas monteilii Pseudomonas putida Pseudomonas stutzeri Pseudomonas sp. Serratia sp. Sinorhizobium fredii Sphingobium sp. Sphingomonas bisphenolicum Sphingomonas yanoikuyae Sphingomonas sp. Sphingopyxis sp. Streptomyces sp. Fungia Bjerkandera adusta Dichomitus squalens Irpex lacteus Phanerochaete chrysosporium Phanerochaete magnolia Pleurotus ostreatus Pycnoporus cinnabarinus Trametes versicolor

Initial BPA concentration (mg L-1)

First order rate constant (d-1)

Reference

3 – 10 0.3 11 10 – 50 5 – 30 250 – 1,000 0.3 0.3 0.137 250 200 11 0.3 200 0.3 114 0.23 – 1.14 0.3 0.05 10 100 – 1,000 250 250 0.3 250 – 1,000 50 0.3 250 1 60 100 300 100 1 1

0.05 – 0.01b 0.4 0.05 11.4 – 0.01b 0.7 – 0.3b 0.2 – 0.1b 0.4 – 1 0.3 0.005 1.2 0.05 – 0.2 0.1 – 0.2 0.5 – 0.6 0.1 0.2 – 0.4 11.5 4.9 – 6.3 0.4 0.1 – 1.1 0.08 – 1.1 82.9 – 9.1b 1.3 1.5 0.04 – 0.4 0.3 – 0.1b 0.8 – 2.1 0.4 0.8 4.1 7.1 0.1 2.9 0.1 – 1.2 4.3 0.3

202 56 116 211 138 209 56 56 65 198 204 116 56 204 56 201 51 56 208 207 206 198 198 56 209 227 56 198 52 200 55 50 57 52 212

10 10 10 10 10 10 90.9 10 10

0.5 0.2 0.6 0.2 0.2 0.2 0.1 0.03 0.2

85 85 85 85 85 85 89 85 85

9.1 2

0.4 0.2

97 99

c

Algae Chlorella fusca Scenedesmus acutus


38


Page 40 of 44

c

Plantae Caragana chamlagu Eichhornia crassipes Ipomoea aquatica Portulaca oleracea Oryza sativa Rumex crispus Photo-degradation simulated lake water NO3--induced degradation Reactive mineral MnO2 a-FeOOH ZVI

d

97 10 5 11.4 11.4 5.5 40

3.5 d 6.9 d 1.0 d 3.5 d 6.9 d 1.4 d 0.5

117 113 104 110 114 109 112

2 10

2.8 e 1.2 – 30.5

145 146

10 1 60 2

144 – 1,411 22g 1.6h 0.2i

f

168 165 175 160

1076

a

1077

b

1078

c

1079

(k) were calculated assuming pseudo first order kinetics using the following equation: k = – ln(A/A0) / t;

1080

d

1081

e

1082

f

1083

g

[MnO2] = 70 mg L-1 (nominal concentration at pH 7.5). Extrapolated from data shown in Figure 2 of reference 165;

1084

h

[a-FeOOH ] = 100 g L-1 (nominal concentration);

1085

i

Studies performed using purified enzymes are not included; Rate constants decreased with increasing BPA concentrations;

BPA removal occurred by a combination of adsorption and transformation. For comparison purposes, rate constants

When A = 0, A/A0 was assumed to be 0.001;

[NO3-] = 0.16 – 9.6 mM;

[MnO2] = 17.4 – 174 mg L-1 (nominal concentration at pH 7);

[ZVI] = 250 g L-1 (nominal concentration)

1086


39

Page 41 of 44


1087

1088 1089 1090

Figure 1. Major BPA sources (red), sinks and natural attenuation processes in the environment.

1091

AMO: ammonia monooxygenase, HCA: 4-hydroxycumyl alcohol, ROS: reactive oxygen

1092

species, WWTP: wastewater treatment plant.

1093


40


Page 42 of 44

1094 1095 1096

Figure 2. Summary of known pathways and intermediates of BPA degradation/transformation

1097

processes mediated by bacteria, fungi, plants and reactive mineral phases demonstrated in the

1098

laboratory. Only the major intermediates are shown and several other, presumably minor

1099

intermediates that have been reported, are not depicted. IPP: 4-isopropenylphenol; HBAL: 4-

1100

hydroxybenzaldehyde; HBA: 4-hydroxybenzoate; HAP: 4-hydroxyacetophenone; HCA: 4-

1101

hydroxycumyl alcohol; HQ: hydroquinone; MME/DME: monomethyl/dimethyl ether.


41

Page 43 of 44


Figure1 167x98mm (144 x 144 DPI)



Figure2 633x456mm (72 x 72 DPI)


Page 44 of 44

Fate of Bisphenol A in Terrestrial and Aquatic Environments

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