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Metal-organic frameworks (MOFs) are composed by metal ions (or metal clusters) and bridging organic ligands, which have shown great application potential in biomedicine and environment fields. However, MOFs would inevitably be transformed or degraded in the environment and/or organisms during applications, which could change their structures and limit the applications. At the same time, MOFs could bring potential risks to human and environmental health. Therefore, the effects of simulated biological physiological conditions (such as pH, ions, protein/enzyme, microorganism) and environmental conditions (such as moisture, redox substances, light, temperature) on the degradation of MOFs are systematically summarized in this review, and the related degradation mechanisms are discussed. In the human digestive system and blood, pH and PO43− are the main factors that affect the degradation of MOFs, respectively, which could lead to the degradation of MOFs by directly destroying coordination bonds in MOFs and acting on metal centers in MOFs. While in the environment, water molecules are the key factors that affect the stability of MOFs, which could replace the organic ligands in MOFs through substitution reaction or destroy the coordination bonds in MOFs through hydrolysis, resulting in the degradation of MOFs. This review provides support for clarifying the degradation behaviors of MOFs and the impact of degradation products on human health and environmental fate. Finally, combining with existing issues, this article highlights the prospects and suggestions for future research perspectives of MOFs degradation, providing a theoretical basis for the green and sustainable application of MOFs.

Due to excellent properties such as ease of processing, transportation and storage, as well as outstanding chemical stability, plastics are widely used in daily production and life, which induce the possibility of plastics entering the natural environment and long-term staying in the environment. Plastics could release into different environmental compartments such as sea, lake, soil and sediment through various ways, and have toxicity to organisms through ingestion and bioaccumulation. In particular, the physical and chemical properties of plastics can be changed by biotic and abiotic factors in the environments. Biodegradation is one of the inevitable environmental processes of plastics, and it is also an eco-friendly treatment for plastic waste. Therefore, exploring the biodegradation of plastics has great practical and ecological significances. In this review, the biodegradation process of plastics by animals, plants, microorganisms and enzymes were summarized and analyzed, and the biodegradation mechanisms of plastics were revealed. Animals (e.g., insects) could physically crush plastics by gastric grinding and chewing, and also change the chemical properties of the ingested plastics under the complex biochemical reactions in their bodies. Current researches on the biodegradation of plastics by animals mainly focus on insects and their intestinal microorganisms. Limited investigations focused on plants, in which the colonization of plants (e.g., algae) on the plastics surface was observed. The biodegradation by plants could be attributed to the enzymes existing on the surface of stems and leaves or secreting by roots. Therefore, it can be concluded that the biodegradation of plastics by animals and plants is closely related to microorganisms and enzymes. It is noted that the microorganisms from animals, plants and external environments showed similar pathways on plastics biodegradation. The microorganisms could first colonize on the plastics surface to form a biofilm, and then secrete extracellular or intracellular enzymes to contact with the plastics and decompose the plastics into low molecular weight oligomers, dimers or monomers through catalysis of their binding sites. These generated oligomers, dimers and monomers can be absorbed by microorganisms as a carbon source through in vivo metabolic mechanism (e.g., the assimilation of organisms). Finally, plastics can be mineralized to CO2, CH4 and/or H2O. In addition, this review deeply analyzes the factors that affect the efficiency of plastic biodegradation, including plastic properties (e.g., plastics types, glass transition temperature, surface hydrophobicity, types of additives), organism species (e.g., insects, marine animals, plants, bacteria, fungi) and abiotic environmental factors (e.g., temperature, oxygen content, solar radiation, physical abrasion). In addition, the environmental behavior of the products after plastics biodegradation was discussed. The small-sized fragments, oligomers and plasticizers released after the biodegradation of plastics could further show different effects on organisms and ecosystems. In conclusion, challenges and future research perspectives for the plastic biodegradation were suggested at the end of this review. Further investigation on the biodegradation mechanisms of plastics and the degradation and fate of plastics additives are suggested. More efforts are recommended on the environmental behaviors and possible biological impacts of the degraded microplastics and nanoplastics. The unified measurement standards and characterization methods should be explored. This review will provide forceful scientific supports for better understanding the environmental fate of plastics.

Microplastics (MPs) are recognized as vectors for the transport of organic contaminants in aquatic environments in addition to their own adverse effects on aquatic organisms. Per- and polyfluoroalkyl substances (PFASs) are widely present in aquatic environments due to their widespread applications, and thus coexist with MPs. Therefore, we focus on the interaction of MPs and PFASs and related combined toxicity in aquatic environments in this work. The adsorption of PFASs on MPs is critically reviewed, and new mechanisms such as halogen bonding, π-π interaction, cation-π interactions, and micelle formation are proposed. Moreover, the effect of MPs on the transport and transformation of PFASs in aquatic environments is discussed. Based on four typical aquatic organisms (shellfish, Daphnia, algae, and fish), the toxicity of MPs and/or PFASs at the organismal or molecular levels is also evaluated and summarized. Finally, challenges and research perspectives are proposed, and the roles of the shapes and aging process of MPs on PFAS biogeochemical processes and toxicity, especially on PFAS substitutes, are recommended for further investigation. This review provides a better understanding of the interactions and toxic effects of coexisting MPs and PFASs in aquatic environments.

Background: Airborne nanoparticles can be inhaled and deposit in human alveoli, where pulmonary surfactant (PS) molecules line at the alveolar air-water interface to act as the first barrier against inhaled nanoparticles entering the body. Although considerable efforts have been made to elucidate the mechanisms underlying nanoparticle-PS interactions, our understanding on this important issue is limited due to the high complexity of the atmosphere, in which nanoparticles are believed to experience transformations that remarkably change the nanoparticles’ surface properties and states. By contrast with bare nanoparticles that have been extensively studied, relatively little is known about the interactions between PS and inhaled nanoparticles which already adsorb contaminants. In this combined experimental and computational effort, we investigate the joint interactions between PS and graphene with coexisting benzo[a]pyrene (BaP).Results: Depending on the BaP concentration and molecular agglomeration, different nanocomposite structures are formed via BaPs adsorption on graphene. Upon deposition of graphene carrying BaPs at the pulmonary surfactant (PS) layer, competition of interactions between different components determines the interfacial processes including BaP solubilization, graphene translocation and PS perturbation. Importantly, BaP adsorbed on graphene is solubilized to increase its bioavailability and inhibit the PS biophysical function. Translocation of graphene across the PS layer is facilitated by BaP adsorption through segregating it from contact with PS, while translocation of graphene oxide is suppressed due to increase of the surface hydrophobicity. Graphene extracts PS molecules from the layer, and the resultant PS depletion declines with graphene oxidation and BaP adsorption.Conclusion: Graphene showed high capacity of adsorbing BaPs to form nanocomposites, which were inhaled and deposit in alveoli, where competition of interactions between different components determined the interfacial processes of BaP solubilization, graphene translocation and PS perturbation.

With wide applications, the production of carbon nanomaterials (CNMs) is rapidly increased, and the CNMs in nanoproducts during consumption are inevitably released into the environment. CNMs in the environment can interact with co-existing pollutants (e.g., organic pollutants, heavy metals and other nanoparticles), thus affecting their fate and toxicity. Therefore, the joint toxicity of CNMs and co-existing pollutants cannot be neglected during the environmental risk assessment of CNMs. In this review, we firstly summarized the mechanisms of direct and indirect toxicity of CNMs; then the joint toxicity of CNMs with other pollutants was emphatically discussed to elaborate the difference with single toxicity; finally, the evaluation methods, research level and challenges of the joint toxicity of CNMs with co-existing pollutants were analyzed and prospected. This review will provide a theoretical basis for the accurate assessment and better understanding of the environmental risks of CNMs.

Nanotechnology is developing rapidly and has been widely applied in chemical engineering, electronic engineering, energy sources, environmental protection and other fields. Engineered nanoparticles (ENPs) are inevitably releasing into the aquatic environment during the entire life cycle of nano-products, thus exposing to aquatic organisms, which has been well investigated in the past decade. ENPs could be internalized by different types of aquatic organisms including algae, plants, and fish. Therefore, the possible transfer of ENPs between different aquatic organisms has attracted attentions, and many important publications reported the trophic transfer of ENPs through aquatic food chains. This review first provided the internalization and distribution of ENPs in different aquatic organisms including microorganism, aquatic plants, invertebrate and fish: (1) For microorganisms, ENPs are mainly adsorbed on the cellular surface, and then be internalized, and the internalization could lead to subsequent toxicity; (2) ENPs could be taken up by the roots of aquatic plants, and then be migrated to stem and leaves by evapotranspiration; (3) ENPs were mainly located in the hepatopancreas and gut of invertebrates; (4) ENPs could be accumulated in the gill and digestive tract of fish, and then distributed into other tissues such as liver and spleen. Secondly, the research status of ENPs in aquatic food chain were summarized and analyzed. The transfer of ENPs was observed in both freshwater and marine food chains, and current investigations on freshwater food chains contributed a larger proportion than in the marine. Most of the researches proved that ENPs can transfer along the food chain, but the biomagnification was only reported in a few publications. Thirdly, the influence of water chemistry factors and ENPs properties on the transfer of ENPs via aquatic food chains was discussed. For the water chemistry factors, like light illumination, solution pH, co-existing compounds such as natural organic matter, ionic strength, and oxidation–reduction potential all altered the bioavailability of ENPs and played important roles in the trophic transfer of ENPs. For the properties of ENPs, it is observed that: (1) the metal ions which were released from metal-containing ENPs can improve the bioavailability of ENPs in aquatic organisms; (2) different coatings of ENPs brought different or opposite surface charge on the surface, thus changing the internalization and bioavailability of ENPs. For example, ENPs with positively charged surface could be easily adsorbed on the surface of aquatic organisms than the negatively charged particles; (3) different size and shape influenced the contacting area between the ENPs and organisms, and ENPs with smaller size were usually more readily enter aquatic organisms, which were deeply discussed in this review. Finally, the current researches, prospects and challenges on ENPs transfer in the aquatic food chains are addressed.

Since industrial revolution, anthropogenic carbon dioxides releasing into the atmosphere are absorbed by the global oceans, which leads to measurable decrease of surface seawater pH and carbonate ion concentration ([CO32−]), thus causing the acidification of global ocean. Besides direct threatening the stability of marine ecosystems, ocean acidification (OA) could also change the environmental processes of marine pollutants, thus indirectly influencing their toxicity to marine organisms. By selecting heavy metals and metallic nanoparticles (MNPs) as the model marine pollutants, this review deeply summarizes the dominant mechanisms on species changes of heavy metals, and the dissolution, dispersion and transport of MNPs based on the causes of OA. The effects of OA-induced environmental process alteration on the toxicity of heavy metal and MNPs to marine organisms at different levels are explored. The differences among the observed toxicological data are analyzed. Finally, the key research challenges and future directions on the toxicity of OA and coexisting pollutants are provided.

The methods of life cycle assessment and cost-effectiveness analysis were employed to compare the energetic, environmental, and economic performances of three household-based modern bioenergy (MBE) utilization systems: biogas utilization (BGs), biomass briquette utilization for energy only (BBs), and biomass briquette utilization for energy and biochar (BCs). The results showed that all the three MBEs performed positive results in net energy output, which ranged from 0.86 to 3.75 MJ (kg crop residues)−1. The positive greenhouse gas reductions also can be acquired when using these three MBEs to substitute for traditional energy, respectively. The cost-effectiveness values of three MBE systems ranged from 343 RMB (tCO2e)−1 for BGs to 598 RMB (tCO2e)−1 for BCs (the CO2e indicates the CO2 equivalent). Considering the supply index analysis results furtherly, the BGs is regarded as the best strategy for the development of MBE in rural households. However, BG leakage during BG digester operations should be received much greater attention to avoid adverse impacts on GHG mitigation.

Crop residue (CR) based-technology has several benefits, including renewable energy production and improvement in the environment and waste biomass management. However, the quantitative assessment of different CR based-technologies in terms of these benefits is limited. In this study, three typical CR-based technologies, CR biochar (CR-BC), CR direct-fired power (CR-DFP), and CR briquette biofuel (CR-BB), were assessed in terms of energy- and environment-related performances using a mixed-unit input-output life cycle assessment method. The results indicated that CR-BB performed better in energy production and air pollution mitigation than the other two technologies. Energy conversion efficiency was suggested as a key factor in determining the potentials of bioenergy production and environmental improvement. Furthermore, based on the energy demand from the Chinese agricultural sector and CR supply in 2012, the energy supply capacity (ESI) of CR-BC, CR-DFP, and CR-BB was estimated to be 24.3, 2.89, and 53.0, respectively, and their total greenhouse gas (GHG) reduction potential (TRGHG) was estimated to be 1.41 × 106, 4.81 × 106, and 1.68 × 106 t CO2e (CO2 equivalent), respectively. Overall, the CR-DFP and CR-BC are recommended for the high-value utilization of CR resources. The findings of this study could provide a basis for the development of CR-based technologies and CR management policy.

Figure S1. Physicochemical characterization of AG-QDs. Figure S2. The hydrodynamic diameter of AG-QDs in DI water and cell culture medium. Figure S3. Cell viability of macrophages after exposure to AG-QDs at different concentrations. Figure S4. Viability of MC3T3-E1 cells after exposure to AG-QDs at different concentrations. Figure S5. Apoptosis of NR8383 cells during 24- and 48-h AG-QDs exposure. Figure S6. Export of AG-QDs by macrophages after AG-QDs internalization. Figure S7. Distribution of internalized AG-QDs in macrophages after exocytosis for 24 and 48 h. Figure S8. Confocal images of AG-QDs distributed in mitochondria, endo-lysosomes, and endoplasmic reticulum after incubation for 24 h. Figure S9. The effect of AG-QDs on the membrane stability of endo-lysosomes. Figure S10. Uptake of GO-QDs by NR8383 cells under confocal imaging. Figure S11. Confocal images of AG-QDs distributed in the cellular cytoplasm under 438 nm fluorescence emission. Figure S12. Distribution of AG-QDs in the cellular cytoplasm after incubation for 12 h. Figure S13. Relative expression of Kapβ2 and Nup98 after exposure to AG-QDs (200 μg/mL) for 12 and 24 h. Figure S14. The high content screening (HCS) images of NR8383 cells after AG-QDs (200 μg/mL) exposure in absence and presence of Hoechst33342 at 347 nm excitation and 483 nm emission. Figure S15. Raman images of cells after exposure to AG-QDs (200 μg/mL) for 24 and 48 h. Figure S16. ROS levels of AG-QDs (200 μg/mL) alone and NR8383 cells after AG-QDs exposure. Figure S17. AFM image of DNA chains that were directly exposed to AG-QDs-FBS (200 μg/mL) for 24 h. Figure S18. Ten representative structural models of AG-QDs. Figure S19. π-π Interactions between the AG-QDs (Structures 7–10) and DNA. Table S1. The number of π bonds between AG-QDs (Structures 1–10) and DNA as obtained by molecular docking. Table S2. The expression of genes in the caspase family after AG-QDs exposure. The macrophages were exposed to AG-QDs for 24 h prior to analysis. Table S3. Sequences of the gene-specific primers used in the quantitative real-time PCR (qRT-PCR) experiment. (DOCX 2467 kb)

We systematically investigated the toxicity mechanism of three graphene-family materials (GFMs), graphene oxide (GO), reduced graphene oxide (rGO) and multi-layer graphene (MG), to algae (Chlorella pyrenoidosa). GFMs exhibited much higher toxicity than other carbon materials (carbon nanotube and graphite), with the 96 h median effective concentration (EC50) values of 37.3 (GO), 34.0 (rGO), and 62.2 (MG) mg/L. Shading effect contributed approximately 16.4% of growth inhibition by GO due to its higher dispersibility and transformation while the other GFMs did not show any shading effect. Hydrophobic rGO and MG more readily heteroagglomerated with algae than GO, thus likely leading to more direct contacts with algae. Flow cytometry results showed significant decrease of membrane integrity after GFM exposure, and rGO caused the highest membrane damage, which was confirmed by the increased DNA and K+ efflux. The observed membrane damage was caused by a combination of oxidative stress and physical penetration/extraction. Moreover, all the three GFMs could adsorb macronutrients (N, P, Mg, and Ca) from the algal medium, thus leading to nutrient depletion-induced indirect toxicity. GO showed the highest nutrient depletion (53% of total toxicity) due to its abundant functional groups. The information provided in this work will be useful for understanding toxicity mechanism and environmental risk of different GFMs in aquatic environments.

This study investigated the adsorption of cadmium (Cd(II)) by cerium oxide nanoparticles (CeO2 NPs) and oyster shells in seawater. The results showed that the addition of Cd(II) significantly inhibited the agglomeration of CeO2 NPs both in DI water and seawater, increased the positive charges of CeO2 NPs in DI water and neutralized the negative charges of CeO2 NPs in seawater. Additionally, CeO2 NPs could adsorb Cd and the bioavailability of Cd was reduced in the presence of oyster shells. This study demonstrated that the adsorption of metals on shells should not be neglected for the accumulation of metals by shellfish.