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Purpose and Scope The Journal of Hazardous Materials publishes full-length research articles, reviews, case studies, and briefs that improve our understanding of the hazards and risks that certain materials pose to humans and the environment, or address ways to control those hazards and related risks. The editors are especially happy to receive proposals for peer-reviewed articles. In order to limit the scope, the following areas are excluded: standard urban processes and waste and wastewater treatment processes; production of explosives. The journal publishes articles with a high impact on: • Characterization of harmful effects of chemicals and materials • Measurement and monitoring of hazardous materials • Transport and fate of HM in the environment • Risk assessment and handling • Physico-chemical and separation processes • Biological processes • Advanced oxidation processes • Safer and cleaner technologies The journal publishes articles with significant innovations and scientific impact. The editors reserve the right to reject, without external review, articles that do not meet these criteria, including articles that: • too closely resemble previous publications, with changed target substrates, materials used, analyzed sites or experimental methods, • optimization of parameters of known processes without new concepts and /or interpretations, • Report on environmental analysis and monitoring of specific geographic areas without presenting new knowledge and/or testing hypotheses, or • Not focusing on the ecological relevance and importance of investigated systems or materials. Editors Professor Diana S. Aga, Department of Chemistry, State University of New York (SUNY) at Buffalo, 359 Natural Sciences Complex, Buffalo, NY 142603000, USA, Tel: (716) 645-4220, E-mail:[email protected]Professor Wonyong Choi, School of Environmental Science and Engineering Pohang University of Science and Technology (POSTECH) Pohang, Korea 790-784 Phone: +82-54-279-2283 Phone: 82-54-279-8299 Email:[email protected]Professor Andrew Daugulis, Department of Chemical Engineering, Queen's University, Dupuis High, Kingston, K7L 3N6, Canada, Tel: (613) 533-2784; Fax: (613) 533-6637; E-MAIL:[email protected]Professor Gianluca Li Puma, Institute of Chemical Engineering, Loughborough University, Loughborough LE11 3TU, UK. Phone: +44 (0) 1509 222 510; Fax: +44 (0) 1509 223 923; E-MAIL:[email protected]Professor Gerasimos Lymberatos, Faculty of Chemical Engineering, National Technical University of Athens, Zografou Campus, Athens 15780, Greece. Phone: +302107723256; Fax: +302107723155; E-MAIL:[email protected]Professor Tay Joo Hwa, Department of Environmental and Water Resources Engineering, School of Civil and Environmental Engineering, Nanyang Technological University, Nanyang Avenue, Singapore 639798, Tel: +65 6790 6228, Fax: +65 6791 Email: 0676,[email protected]Contributing Editors Jaesang Lee, Aquatic Environment Center, National Agenda Research Division, Korea Institute of Science and Technology (KIST), 39-1 Hawolgok-Dong, Seongbuk-Gu, 136-791 Seoul, Sydkorea, Tel.: +82- 2- 958-6947, E-mail:[email protected]Konstantina Marazioti, Laboratory of Biochemical and Environmental Technology, Department of Chemical Engineering, University of Patras, Patras, Greece, Tel.: +30 6976 128157, Email:[email protected]Redaktionsråd R. Andreozzi (Napoli, Italy) J. Baeyens (Tremelo, Belgium) A. Ballester (Madrid, Spain) E. Brillas (Barcelona, Spain) V. Cozzani (Bologna, Italy) Mexico D. Dong (Changchun) , China) ) M. Erdem (Elazig ̆, Turkey) S. Esplugas (Barcelona, Spain) M. Fingas (Alberta, Canada) VK Gupta (Roorkee, India) O. Hamdaoui (Annaba, Algeria)

W.-J. Lee (Tainan, Tajvan, ROC) P. Lianos (Patra, Greece) E. Lima (Porto Alegre, Brazil) T. Mangialardi (Rim, Italy) A. Mittal (Bhopal, India) D. Mohan (New Delhi, India) C. Mulligan (Montreal, Canada) A. Nzihou (Albi, Francisca) J.M. Park (Pohang-si, Republic of Korea) H.J. Pasman (Delft, South Africa) S. Pavlostathis (Atlanta, Georgia, SAD) H. Potgieter (WITS, South Africa)

P. Randall (Cincinnati, OH, SAD) K. Reddy (Chicago, IL, SAD) J. Sneddon (Lake Charles, Louisiana, SAD) M. Soylak (Kaiser, Turska) M. Talawar (The Hague, Nizozemska) Townsend ( Gainesville). , FL, SAD) S. Tunali Akar (Skis ̧ehir, Turkey) S. Wang (Perth, WA, Australia) H.-Q. Yu (Hefei, China) J. Yu (Wuhan, China) S. Zeman (Pardubice, Tzekkiet) A.I. Zoumboulis (Solun, Greece)

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Journal of Hazardous Materials (ISSN 0304-3894). Volumes 182-192 are planned to be published in 2011. Subscription prices are available on request from the publisher or from the Elsevier customer service nearest you, or from this journal's website (http://www.elsevier.com/locate/jhazmat). More information about this journal and other Elsevier products is available through the Elsevier website (http://www.elsevier.com). Subscriptions are accepted only on a prepaid basis and are registered on a calendar year basis. Items are sent by regular mail (country within Europe, air delivery outside Europe). Priority pricing is available upon request. Requests for missing issues must be made within six months of the date of mailing. Orders, complaints and diary questions: contact the Elsevier customer service nearest you: St. Louis: Elsevier Customer Service, 3251 Riverport Lane, Maryland Heights, MO 63043, USA. phone: (877) 8397126 [toll free in the US]. (+1) (314) 4478878 [outside US]; fax: (+1) (314) 4478077; EMAIL:[email protected]Oxford: Elsevier Customer Service, The Boulevard, Langford Lane, Kidlington, Oxford OX5 1GB, UK; telefon: (+44) (1865) 843434; faks: (+44) (1865) 843970; EMAIL:[email protected]Tokio: Elsevier customer service department, 4F Higashi-Azabu, 1-Chome Bldg, 1-9-15 Higashi-Azabu, Minato-ku, Tokio 106-0044, Japan, telephone: (+81) (3) 5561 5037; (+81) (3) 5561 5047, e-mail:[email protected]Singapore: Elsevier Customer Service, 3 Killiney Road, #08-01 Winsland House I, Singapore 239519; phone: (+65) 63490222; fax: (+65) 67331510; E-MAIL:[email protected]doi:10.1016/S0304-3894(11)01319-7

Journal of Hazardous Materials 196 (2011) 1-15

Content is available on SciVerse ScienceDirect

Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Review

Effects of the presence of sulfonamides in the environment and their impact on human health b ´ Wojciech Baran a, Ewa Adamek a,*, Justyna Ziemianska, Andrzej Sobczak a,b a b

Silesian Medical University, Department of General and Analytical Chemistry, Jagiello´nska 4, 41-200 Sosnowiec, Poland Department of Occupational and Environmental Health, Ko´scielna 13, 41-200 Sosnowiec, Poland

article

i n f o

Article history: Received March 10, 2011 Received in revised form July 22, 2011 Accepted August 31, 2011 Available online September 6, 2011 Keywords: Sulfonamides Biotransformation Ecotoxicity Environmental risk Drug resistance

a b s t r a c t Global production and consumption of pharmaceutical products is constantly increasing. Anti-infectives are particularly important in the modern treatment of microbial infections. Sulfonamides have been widely used for a long time as anti-infectives and are still widely used today. This review presents the most common types of sulfonamides used in health and veterinary medicine and discusses the problems associated with their presence in the biosphere. Based on the analysis of more than 160 documents, it was determined that small amounts of sulfonamides in the environment originate mainly from agricultural activities. These drugs have caused changes in the population of microbes that can be potentially dangerous to human health. This risk to human health may be global in scope, and management activities have not been effective in reducing the risk. © 2011 Elsevier B.V. All rights reserved.

Contents 1. 2. 3. 4. 5. 6. 7. 8.

9. 10. 11. 12.

Introduction. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Physico-chemical properties of SN. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Mechanism of antibacterial action of SN. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Use of SNs. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Estimated MV consumption. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Occurrence of SN in the environment and food. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Ecotoxicity of SN. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . SN degradation in organisms and the environment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.1. Metabolism of SN in mammals. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.2. Biodegradability of SN. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . 8.3. Physico-chemical methods of decomposition of SN. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Removal of SN from wastewater. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Environmental risk assessment. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Creates drug resistance. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Findings. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Thank you. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . Bibliographic references. . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . . .

1. Introduction The global production and consumption of pharmaceutical products is constantly growing at a rate higher than the global rate

1 2 2 3 4 5 6 7 7 7 8 8 8 9 12 12 12

population growth. After use, large quantities of drugs are released into the environment in the form of human and animal feces and unused waste [1]. The persistence of drugs in the environment, their speed of dispersion and ability to accumulate in the biosphere vary. But their high biological activity indicates that these drugs, even in trace amounts, can cause significant changes in the biosphere. An example of such changes in the last decade of the 20th century is

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

Fourth

H2N

Lille

N1 R

THE

Fig. 2. Chemical structure of SN with bacteriostatic properties.

• excessive amounts of SN are introduced into the biosphere (common practice is illegal and/or uncontrolled application of SN to healthy livestock), • local concentration of SN in the environment and the risk associated with this problem is very high, • SN can remain in the environment for a long time. 2. Physico-chemical properties of SN Sl. 1. Possible fates of SNs residues and resistance genes (SNsR) in the environment.

the phenomenon of feminization of fish with sex hormones caused by anthropogenic pollution of European rivers [2]. This is why pharmaceutical products are classified as particularly dangerous environmental pollutants. As a result, research and multinational projects (eg REMPHARMAWATER [3], POSEJDON [4], KNAPPE [5], ERAPHARM [6] and ECO-SENS [7]) have been carried out to find answers to the following questions: • Which medicines pose the greatest risk to the environment? • How can we effectively control the quantities and effects of substances on the environment? • How can we successfully reduce their emissions into the environment? Antibiotics are a group of medicinal substances with an environmental impact that can be particularly harmful to human health. Unfortunately, their frequency in environmental samples is very high [1,5,8-18]. Historically, sulfonamides (SN) have been used for the longest time as synthetic antibiotics. Recently, large amounts of SN are used in animal husbandry, especially as veterinary drugs. Based on these drugs, we can achieve a reliable assessment of the effects and consequences of long-term use of anti-infectives on human health and the environment. The report of the State Office for Nature, Environment and Consumer Protection in North Rhine-Westphalia (Germany) published in 2007 presents a review of the literature on the effects of introducing SN into the environment [1]. In most published articles, authors almost exclusively assessed the risk of SNs based on their use, toxicity, and environmental removal efficiency. The evidence presented in this context led to the conclusion that the presence of substances in the environment is an insignificant problem in terms of quality of life. However, most articles did not consider the effect of anti-infective agents on the development of microbial drug resistance. The impact of antibiotics that appear in the environment on the creation and spread of drug-resistant microorganisms is important from the point of view of human health (Figure 1). This influence is much more widespread in recent decades due to the process of globalization. The purpose of the work is to show that:

Since the early 1940s, more than 150 SNs, sulfonamide derivatives, have been used in medicine and veterinary medicine as antibacterial drugs [19]. The type of structure shown in Fig. 2 corresponds to synthetic antimicrobial agents containing a sulfonamide group. Such a molecule should have a free amino group (–N4H2) at one end. SNs are a group of synthetic bacteriostatic drugs classified according to the Anatomical Therapeutic Chemical (ATC) classification index in the group of antibacterial drugs for systemic use (subgroup J01E) [20]. Many derivatives of SN have also been used as antiprotozoal agents [21] and herbicides [19], and complexes of SN with Ag+ and Zn2+ have been used as antifungal agents [22]. In addition, SNs were the most commonly used components of several combination drugs with trimethoprim (TMP). The characteristics of commonly used SNs are shown in Table 1. SNs are polar molecules with amphoteric properties. Their amino nitrogen (N4) is protonated at pH 2-3, while the amide nitrogen (N1) is deprotonated at pH 4.5-11 [10,23]. The SN presented in this text are small molecules (molecular weight 177–300 g mol−1), are soluble in water (with the exception of SGM and sulfasalazine) and have a low Henry's constant (1.3 × 10−12 -1.8) × 10 −8) values [9,10,24]. They are easily adsorbed from the soil (values of the soil distribution coefficient are 0.6–7.4 l kg−1) [9]. Therefore, these SNs spread easily and quickly in the environment, but their properties should limit their accumulation in defined habitats. SNs are not easily adsorbed on activated carbon [1,4]. They are classified as photo- and heat-stable substances with a half-life of decomposition (DT50) >1 year [24]. They can undergo alkaline hydrolysis and coupling reactions with phenols and amines and easily react with the hydroxyl group HO• [10,11,25]. 3. Mechanism of antibacterial action of SN As shown in Fig. 3, antibacterial SN act as competitive inhibitors of the enzyme dihydropteroate synthase (DHPS), which catalyzes the conversion of para-aminobenzoate (PABA) to dihydropteroate (AHHMD), the precursor of leaves. synthesis. Tetrahydrofolic acid (THF) participates in the synthesis of nucleic acids, which are necessary as building blocks of DNA and RNA. The mechanism of action of herbicide SN is similar. As a result, it is possible to inhibit the synthesis of nucleic acids and, consequently, proteins [18,27]. SNs also inhibit the permeability of the bacterial cell wall to glutamic acid, which is also an essential component in the synthesis of folic acid. However, SNs do not inhibit the growth of microorganisms that:

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

Table 1 Common names, CAS number and structure of selected SNs. Common name for SN

CAS number

ATC Classification Index

Abbreviation

– R

Sulfanilamid Sulfacetamid Sulfacarbamid Asulam (ζιζανιοκτόνο) Karbutamid Sulfatiourinstof Sulfaguanidin

63-74-1 144-80-9 547-44-4 3337-71-1 339-43-5 515-49-1 57-67-0

J01EB06, D06BA05, QJ01EQ06 S01AB04 J01EC20 (sa SDZ i SDM) – A10BB06 J01EB08 A07AB03

SAD SCT SC

–H –COCH3 –CONH2 –COOCH3 –CONH(CH2)3 CH3 –CSNH2 C(NH2)2

sulfatiazol

72-14-0

D06BA02, J01EB07, QJ01EQ07

STZ

THIS SGM

S N

H3C sulfafurazol, sulfizoksazol

127-69-5

J01EB05, S01AB02, QJ01EQ05

CH3

N sulfamethoxazole

723-46-6

J01EC01, QJ01EQ11

THE

SMX

CH3

sulfamoksol

Sulfapyridine

729-99-7

144-83-2

J01EC03

SMM

J01EB04, QJ01EQ04

SPION

THE

CH3

N N

Sulfadiazine

68-35-9

J01EC02, QJ01EQ10

SDZ

N sulfamethox, sulfamethoxydiazine

651-06-9

J01ED04

I 3

WE ARE

N sulfamerazin

127-79-7

J01ED07

SMR

N CH3 CH3

N Sulfametazin, Sulfadimidin

57-68-1

J01EB03, QJ01EQ03, QP51AG01

SDM

NCH3OCH3

Sulfadimethoksin

122-11-2

J01ED02, QJ01EQ09, QP51AG02

SDT

N N OG3 N N

Sulfamethoxypyridazine

80-35-3

Sulfachlorpyridazine

80-32-0

J01ED05, QJ01EQ15

I 3

SMP

N N SCP

N N Sulfadoxine

2447-57-6

QJ01EQ13

SDX

H3CO

• need the presence of folic acid in the environment, • have a high concentration of PABA, or • have altered metabolic pathways (drug resistance). 4. Use of SNs SNs are active against a wide range of gram-positive and many gram-negative bacteria, including species from the genus

OCH3

Streptococcus, Staphylococcus, Escherichia, Neisseria, Shigella, Salmonella, Nocardia, Chlamydia and Clostridium. In addition, SNs have been used against protozoa (eg Toxoplasma gondii), parasites (eg Plasmodium malariae) and fungi (eg Pneumocystis carinii). SMX, SCT or sulfasalazine belong to SNs commonly used in medicine, while SDM, SDT, SMR, SDZ, STZ are more commonly used in veterinary medicine (different SNs were used in different countries). In addition, SN was added to animal feed

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

H2N N 2-amino-4-hydroxy-6-hydroxymethyl7,8 dihydropteridine diphosphate

H N

H H

CH2

H2N

THE

H N

N N

P ~ P

ATP AMP

H H CH2

OH (2-amino-4-hydroxy-6-hydroxymethyl-7,8-dihydropteridine

(AHHMD)

(AHHMP) dihydropteroacintetaza (DHPS) EC 2.5.1.15 - P-P og H2O

SO2NHR

H2N

Sulfonamides

COOH

H2N

4-aminobenzoesire (PABA)

H N

H2N N

H H

Dihydropteroinsir COOH

CH2HN

N OH

CH2

COOH

H2N CH

Glutaminsyre

ATP

COOH H N

H2N N

H H

THE

CH2

CH2HN

C HN

CH2

COOH [O] [H]

COOH

Dihydrofolic acid (DHF) reductase (DHFR) EC 1.5.1.3

Folic acid N-[p-[[(2-amino-4-hydroxy-6-pteridinyl)methyl]amino]benzoyl]-L-glutamic acid

[THE]

H2N N

H N N H

H H H CH2HN

does

CH2

THE

HN CH

COOH

Tetrahidrofolinsir (THF)

COOH

Fig. 3. Scheme for SN pathways, based on Wilson & Gisvold's textbook of organic medicinal and pharmaceutical chemistry [26].

premix for feeding young animals. For example, in Denmark in 2009, consumption of SN with TMP per kg of meat produced was as follows [28]:

However, the use of Asulam can lead to contamination of honey with residual SN [29]. In 2008, it was withdrawn from use in EU countries.

• • • •

5. Estimated use of SNs

pigs 4.82 mg, cattle 17.2 mg, broilers 0.033 mg, farmed fish (aquaculture) 58.5 mg.

In addition, SNs can be used in commercial beekeeping (they protect bees from bacterial diseases, e.g. In agriculture, Asulam sulfonamide is widely used as a herbicide. It is effective against dicotyledonous weeds, e.g.

An accurate estimate of the global consumption of all substances would be difficult, if not impossible. The authors of the KNAPPE project estimate that the global consumption of drugs used in human and veterinary medicine has reached 100,000 tons per year [5]. Based on information from the Union of Concerned Scientists, Sarmach et al. reported that Americans consumed 16,000 tons of antibiotics annually at the beginning of the 21st century [9]. SNs used in veterinary medicine represented about 2.3% of the total amount of antibiotics. In Europe

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

18 16

active compounds (t)

14 12 10 8 6 4 2 0 1990. 1992. 1994. 1996. 1998. 2000. 2002. 2004. 2006. 2008. 2010.

year Fig. 4. Dynamics of SN and TMP consumption in Denmark in the years 1990-2009. [28,31].

countries, this value varies from 11 to 23% [9]. According to other authors, the global consumption of antibiotics (drugs against infections) varied from 100,000 to 200,000 tons per year, including 50-75% used in veterinary medicine and animal husbandry [1,24]. It was likely that more than 20,000 tons of SN with bacteriostatic properties were introduced into the biosphere each year (excluding drugs introduced as herbicides). Since the end of the 20th century, Scandinavia and other countries in Europe and North America have introduced restrictions on the use of antibiotics (including SN) in animal husbandry. The use of antibiotics as growth factors in livestock farming in the EU has been banned since January 1, 2006 [30]. However, reports on the consumption of pharmaceutical products in different countries have not shown a decrease in the use of these drugs. Fig. Fig. 4 shows the dynamics of SN consumption with TMP in Denmark in the years 1990-2009 [28,31]. The decrease in the use of SN in animal husbandry occurred in the mid-1990s, and was associated with the introduction of administrative restrictions regarding the use of these drugs in animal feed. Although the ban is still in force, the use of SN in agriculture is similar to 1994. In our opinion, the plots in Fig. 4 global trends in SN consumption in animal husbandry and medicine. 6. Occurrence of SN in the environment and food The first publication containing quantitative data on the presence of SN in river water was published in 1982 [9]. However, systematic studies of the quantification of SNs in environmental matrices became possible after the development of highly sensitive analytical methods. According to the US Environmental Protection Agency, the detection limit during routine analytical procedures using SPE/HPLC-MS/MS techniques for selected SNs was below 10−9 g l−1 (e.g. for SDT the detection limit was 1 × 10 − 10 g l −1). A detailed account of analytical techniques and detection limits for drugs (including SN) in environmental samples is discussed by García-Galán et al. [18] and Seifrtová et al. [32]. At the detection level described, SNs were detected in 27% of rivers and streams in the United States [11], in almost all surface waters in France and Taiwan [33,34], and in 100% of wastewater samples [13, 35,36]. . According to Vulliet and CrenOlivé, the frequencies of SMX in surface and groundwater in the Rhône Alpes region of France were 37 and 66%, respectively [ 37 ]. In commercially available, Italian natural mineral water frequency

των SNs ήταν 50% (σε 4 από τα 8 δείγματα που ερευνήθηκαν) [38]. GarcíaGalán et al. [36] περιέγραψε λεπτομερώς τη συχνότητα εμφάνισης 19 επιλεγμένων SNs στα λύματα. Επιπλέον, μεταβολίτες των SNs, κυρίως Ν4-ακετυλοσουλφοναμίδια (N4-AcSNs), ταυτοποιήθηκαν επίσης σε περιβαλλοντικά δείγματα [11,39]. Οι συγκεντρώσεις των SN στο περιβάλλον υπέστησαν σημαντικές διακυμάνσεις, οι οποίες εξαρτήθηκαν κυρίως από τον τύπο της μήτρας και τον τύπο του SN [36]. Επιπλέον, τα αποτελέσματα που προέκυψαν μπορεί να εξαρτώνται από τον τόπο δειγματοληψίας, την ημέρα της εβδομάδας [40] και ακόμη και την ώρα της ημέρας [41]. Ωστόσο, ήταν σημαντικό να σημειωθεί ότι τα δεδομένα σχετικά με τον προσδιορισμό των SNs σε περιβαλλοντικά δείγματα θα μπορούσαν να περιέχουν σημαντικά σφάλματα. Η αιτία αυτού μπορεί να είναι η ατέλεια της αναλυτικής διαδικασίας που χρησιμοποιείται και η εσφαλμένη (ημιτελής) εξαγωγή δειγμάτων. Για παράδειγμα, η ανάκτηση SNs από δείγματα εδάφους κυμαινόταν από 5 έως σχεδόν 294%, αλλά οι συγγραφείς βρήκαν ότι η «παρουσιαζόμενη μέθοδος χαρακτηρίζεται από καλή επιλεκτικότητα» [42]. Η αποτελεσματικότητα ανάκτησης εξαρτάται από διάφορες παραμέτρους, συμπεριλαμβανομένων των στρατηγικών εξαγωγής/καθαρισμού [43] και του τύπου της μήτρας [44]. Μια περίληψη της εμφάνισης SN, ανάλογα με τη μήτρα, φαίνεται στο Σχ. 5 και οι μέγιστες τιμές δίνονται στον Πίνακα 2. Τα δεδομένα που παρουσιάζονται βασίζονται στις μέγιστες τιμές που περιγράφονται στη βιβλιογραφία. Οι συγκεντρώσεις SNs στα δείγματα αυξήθηκαν ως εξής: θαλασσινό νερό < υπόγεια ύδατα < επιφανειακά ύδατα < επεξεργασμένα λύματα < ακατέργαστα (ακατέργαστα) δημοτικά λύματα < νοσοκομειακά λύματα < ενεργοποιημένη ιλύς < έδαφος < απορροή από γεωργική γη < στραγγίσματα από χωματερή < κοπριά. Λόγω των χαμηλών συγκεντρώσεων και της χαμηλής αφθονίας των SNs, η παρουσία ιχνοποσοτήτων αυτών των φαρμάκων στο πόσιμο νερό δεν θεωρήθηκε σημαντικό πρόβλημα. Οι μέγιστες συγκεντρώσεις βρέθηκαν σε πρόσφατα αφαιρεμένα κλινοσκεπάσματα [58] και κοπριά από χοίρους που τρέφονταν με δίαιτες που περιείχαν SNs, κυρίως SDM [59]. Αυτό το SN εμφανίστηκε σχεδόν στο 50% των δειγμάτων (η μέση συγκέντρωση του φαρμάκου ήταν 7 mg kg−1). Επιπρόσθετα, άλλα SNs ταυτοποιήθηκαν σε δοκιμασμένα δείγματα (π.χ., για SDZ, η μέγιστη συγκέντρωση ήταν 35,2 mg kg−1). Ευτυχώς, ακόμη και η βραχυπρόθεσμη αποθήκευση κοπριάς θα μπορούσε να οδηγήσει σε σημαντική μείωση της συγκέντρωσης των SNs [58]. Οι υψηλότερες επιτρεπόμενες συγκεντρώσεις SNs στα τρόφιμα καθορίστηκαν σε διοικητικούς κανονισμούς. Η Ευρωπαϊκή Ένωση υιοθέτησε μέγιστη συγκέντρωση SN 100 ␮g kg−1 σε ζωικά τρόφιμα [61]. Στην Πολωνία, η μέγιστη επιτρεπόμενη συγκέντρωση Asulam σε φρούτα και λαχανικά είναι 0,5 mg kg−1 [62]. Η εμφάνιση SNs σε ιστούς εκτρεφόμενων ψαριών ήταν τυχαία, π.χ. στη Σλοβενία, υπολείμματα SNs βρέθηκαν σε 14 από τα 2363 δείγματα [63]. Τα υπολείμματα SNs σε βρώσιμα θαλάσσια τρόφιμα ανιχνεύθηκαν σπάνια, ωστόσο η συγκέντρωση των SNs στον ιστό του κοινού χελιού (Anguilla anguilla) ήταν πάνω από 5 mg kg−1 [64]. Στις χώρες της ΕΕ, η εμφάνιση υπολειμμάτων SNs σε βρώσιμους ιστούς ζώων εκτροφής ήταν ασήμαντη. Σύμφωνα με την «Έκθεση για το 2006 σχετικά με τα αποτελέσματα της παρακολούθησης καταλοίπων σε τρόφιμα ζωικής προέλευσης στα κράτη μέλη», SNs, σε συγκεντρώσεις πάνω από τα μέγιστα επιτρεπόμενα όριά τους, εντοπίστηκαν σε 0,006, 0,05, 0,08, 0,97 και 3,86% των δειγμάτων πουλερικών, βοοειδή, χοίρους, αυγά και κουνέλια, αντίστοιχα [65]. Αν και η χρήση των SNs στη μελισσοκομία απαγορεύεται στην ΕΕ, η συχνότητα αυτών των φαρμάκων σε δείγματα μελιού είναι υψηλή. Στην Πολωνία, έχει υπολογιστεί ότι σχεδόν το 10% των δειγμάτων μελιού περιέχουν υπερβολικές ποσότητες SNs, δηλαδή πάνω από τη μέγιστη επιτρεπόμενη συγκέντρωση. Τα αποτελέσματα που αναφέρουν οι Κινέζοι ερευνητές είναι πολύ λιγότερο αισιόδοξα. Προσδιορίστηκαν υψηλές συγκεντρώσεις SN στα παραπροϊόντα χοίρων (σχεδόν 74 mg kg−1 SDT και 73 mg kg−1 STZ) και στα παραπροϊόντα πουλερικών (46 mg kg−1 SDZ) [66]. Ακόμη πιο ανησυχητικό είναι το γεγονός ότι το 75% των δειγμάτων κρέατος περιείχαν SNs σε συνολική συγκέντρωση >100 ␮g kg−1 [67]. Τα SNs θα μπορούσαν να απορροφηθούν και να συσσωρευτούν από φυτά που γονιμοποιούνται με κοπριά (οι υψηλότερες συγκεντρώσεις SN προσδιορίζονται στις ρίζες και τα φύλλα [9,68-71]. Για παράδειγμα, η μέγιστη συγκέντρωση

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

1000 000 100 000

-1

SN concentration (i l )

10 000 1 000 100 10 1 0,1 0,01

Drink bottled mineral water from underground water

Seawater extract

Waste water / inflow

Hospital wastewater

Sewerage

SDZ

Ac-SDM

Snows

Biosolid soil / mud

SDM

SPION

STZ

SMX

SDZ

SPION

SDZ

SMX

STZ

SDZ

SDT

SDM

SMX

SCP

SDM

SDZ

Surface

SMX

STZ

SDM

SMX

SDT

SDM

SMX

SDT

STZ

SMX

0,0 001

Ac-SDM

0,001

fertilizer

Fig. 5. Display of selected SN according to the matrix.

The SDM determined in maize, tomato and lettuce was 0.1 mg kg-1 [69]. Migliore et al. [71] reported that cosmopolitan weeds (Amaranthus retroflexus and Plantago major) showed a high bioaccumulation tendency. In the tissues of these plants grown in media containing SDT, the accumulation rates were 2314 and 6065 mg kg-1, respectively [71]. In our opinion, although the data presented in this section are based on maximum values, they are probably underestimated. Common practice is the excessive and prophylactic use of antibiotics in animal husbandry and the use of manure as fertilizer. Therefore, local actual concentrations

SN in the biosphere are much larger. Moreover, this phenomenon is difficult to control due to the high mobility of SN in the environment. For these reasons, the excessive use of fertilizers containing SN as fertilizers should be prohibited. 7. Ecotoxicity of SN The toxicity of SN for higher organisms (vertebrates) is not high. According to EU Directive 93/67/EEC, investigated SNs can be classified as non-toxic or harmful [72]. The results described in the literature show that SN are not mutagenic or

Table 2. SN concentrations in the environment. Note

Maximum values

2,1 (0–8,5 [8]) ␮g l (SMX); 0,011 ␮g l (SMX) [37] 0,164 ng l−1 (SDT) [38] 0,047 (0,013-0,080) ␮g l−1 (SMX) [38] 0,80 (0,0099-1,11 №1) SMX) 0,053 (0,0002–0,09148 [45]) ␮g l−1 (SDT) 0,87 (0,015–18 [8]) ␮g l−1 (SMX) 2,26 (0,0108–19, 2 ␮g l−1 [46]) 0,0475 ␮g l−1 (SMX) [47] 379,78 (0,66–703,2 [48]) ␮g l−1 (SCP) 46,58 (0, 05–1340 [33]) ␮g l−1 (SMX)

4 2 11 3 39 12 1 7 31

61,11 (0,0269-500 [49]) ␮g l−1 (SDM)

17.78 (0.3–79.9 [51]) ␮g l−1 (SMX) 1.28 (0.353–2.2 [51]) ␮g l−1 (SDZ) 0.517 (0.00366–6, 0 [53]) ␮g l)−1 ␮g l −1 (SDZ) 0.005–4.27 [50]) ␮g l−1 (STZ)

6 2 30 4

ground

22,56 (0,01–113 [54]) ␮g kg−1 (SMX) 99,1 (1,2–197 [54]) ␮g kg−1 (SPY) 211,6 (0,16– 860 [56]) ␮g kg−1 (SN'er)

6 2 10

fertilizer

27,30 (0,23–167 [1]) mg kg−1 (SDM)

59,07 (35,2–91 [1]) mg kg−1 (SDZ)

8.5 ␮g l−1 (PECc for SMX) [8] 0.080 ␮g l−1 (SMX) [38] 3.461 ␮g l−1 (SCT) [45] 25 ␮g l−1 (all SN) [43] ⮐ 0. −1 (SMX) [47] 703.2 ␮g l−1 (SCP) [48] 1340 ␮g l−1 (SMX, from pharmaceutical production) [33] 1158.68 ␮g l−1 (STZ; agricultural waste water) [50] ␮l −1 (SMX) [52] PEC 92.8 ␮g l−1 (all SN) [51] 6.0 ␮g l−1 (SMX) [53] 4.27 ␮g l−1 1 (STZ, wastewater from agricultural treatment plants) [50] 1 ␮ g kg−1 dwd (SPY) [54] 31 ␮g kg−1 (SDM) [55] 400 ␮g kg−1 (STZ; agricultural soil) [57 ] PEC 860 ␮g kg−1 (SCP; soil pore water estimation) [56] 395,730 mg kg−1 (SDT; in bed – day 0) [58] 167 mg kg−1 (SDM) [ 59] 1600 ␮ g l−1 (all SN) [60]

Matrix Drinking water Bottled mineral water Ground water Surface water Sea water Waste water/flush Inlet/waste water

Wastewater from hospitals (after wastewater treatment)

Sludge (after wastewater treatment)

Average / most described SN −1

Landfill a b c d

It was calculated from the maximum values listed in the tables. Number of documents. Predicted concentration in the environment. Dry masses.

−1

Οι W. Baran et al. / Journal of Hazardous Materials 196 (2011) 1-15

καρκινογόνος (τερατογόνος) δραστηριότητα [73]. Από την άλλη πλευρά, στην έκθεση «Environmentally Classied Pharmaceuticals 2009», τα SNs θεωρήθηκαν ως εξαιρετικά τοξικά φάρμακα [74]. Οι αποκλίσεις μεταξύ αυτών των αναφορών προκύπτουν πιθανώς από διαφορετικά κριτήρια που χρησιμοποιούνται για τον καθορισμό ενός κινδύνου. Η Οδηγία 93/67/ΕΟΚ βασίζεται στον περιβαλλοντικό κίνδυνο που ενέχουν οι φαρμακευτικές ουσίες, ενώ η έκθεση «Environmentally Classied Pharmaceuticals» αξιολογεί τόσο τον περιβαλλοντικό κίνδυνο (με βάση τον οξύ τοξικό κίνδυνο για το υδάτινο περιβάλλον) όσο και επιπλέον την ανθεκτικότητα και τη βιοσυσσώρευση SNs στο περιβάλλον. με βάση τις πληροφορίες που δημοσίευσε η Σουηδική Ένωση Φαρμακευτικής Βιομηχανίας [74]). Το Σχ. 6 απεικονίζει την τοξικότητα του SMX σε επιλεγμένους οργανισμούς δοκιμής. Σημαντικά δεδομένα για την οικοτοξικότητα των SNs συνοψίστηκαν σε άρθρα των García-Galán et al. [18] και Isidori et al. [73]. Τα SN είναι πρακτικά μη τοξικά για τους περισσότερους μικροοργανισμούς που δοκιμάστηκαν [4,18,73,75], συμπεριλαμβανομένων επιλεγμένων στελεχών βακτηρίων, όπως Vibrio fischeri, Enterococcus faecalis, Escherichia coli, Pseudomonas aeruginosa και Staphylococcus aureus. Για παράδειγμα, οι τιμές L(E)C50 που προσδιορίστηκαν χρησιμοποιώντας τη δοκιμή Microtox® (V. fischeri) κυμαίνονταν από 16,9 έως 118,7 mg l−1 (για SMX) έως >1000 mg l−1 (για STZ) [73,76, 77]. Οι ισχυρές βακτηριοστατικές ιδιότητες που προκαλούνται από τα SNs θα μπορούσαν να αλλάξουν σημαντικά τη λειτουργία των μικροοργανισμών που ζουν στο περιβάλλον, για παράδειγμα μια σημαντική μείωση της μικροβιακής τους δραστηριότητας [78]. Επιπλέον, ο αριθμός των λιγότερο ευαίσθητων (ανθεκτικών) στελεχών έχει αυξηθεί και ο αριθμός των στελεχών ευαίσθητων σε SNs έχει μειωθεί. Οι Thiele-Bruhn και Beck έδειξαν ότι η απόρριψη ούρων που περιείχαν ακόμη και χαμηλή συγκέντρωση SPY (0,02 mg kg−1) στο έδαφος είχε ως αποτέλεσμα σημαντική μείωση της μικροβιακής δραστηριότητας [78]. Διαπιστώθηκε ότι, στην περίπτωση του SPY, οι τιμές EC10 για τους οργανισμούς του εδάφους κυμαίνονταν από 0,00014 έως 0,16 mg kg−1 (η δοκιμή μείωσης του μικροβιακού Fe(III) και από 0,0071 έως 0,056 mg kg−1 (το υπόστρωμα που προκαλείται τεστ αναπνοής) [79]. Ωστόσο, οι πιο ευαίσθητοι προσδιορισμοί για την παρουσία SNs είναι οι βιοδείκτες που περιέχουν χλωροφύλλη [9,18,73]. Μια εξαιρετικά τοξική επίδραση του SMX στον Synechococcus leopoliensis (EC50 = 0,0268 mg l−1) περιγράφηκε από τους Ferrari et al. [77]. Στην περίπτωση του SMX, οι συγκεντρώσεις μη παρατηρούμενων επιπτώσεων (NOECs) για τα φύκια (Pseudokirchneriella subcapitata και S. leopoliensis) και την πάπια (Lemna gibba) ήταν 0,090 [77], 0,0059 [77] και 0,01 mg l−1 , αντίστοιχα. Αυτό δείχνει ότι ακόμη και οι χαμηλές συγκεντρώσεις SNs μπορεί να επηρεάσουν σημαντικά την ανάπτυξη και την ανάπτυξη των φυτών. Τα SNs μπορούν να συσσωρευτούν σε διάφορους οργανισμούς στην τροφική αλυσίδα και αυτή η συσσώρευση θα μπορούσε να οδηγήσει σε τοπική αύξηση των τοξικών επιδράσεων που προκαλούνται από αυτά τα φάρμακα [9,10,70,71]. Επιπλέον, οι τοξικές επιδράσεις των SNs και άλλων ρύπων θα μπορούσαν να παρουσιάσουν συνέργεια [11,80,81]. Σε επίπεδα περιβαλλοντικής έκθεσης (τα δείγματα περιείχαν 13 μικρορύπους, συμπεριλαμβανομένου του SMX σε συγκέντρωση 46 ng l−1 ) το μείγμα φαρμάκων ανέστειλε την ανάπτυξη ανθρώπινων εμβρυϊκών κυττάρων HEK293, με το υψηλότερο αποτέλεσμα να παρατηρείται ως μείωση 30% στον πολλαπλασιασμό των κυττάρων σε σύγκριση με τους μάρτυρες [81]. Δεδομένου ότι δεν έχουν γίνει αρκετά εκτεταμένα πειράματα σε ασθενείς με μία μόνο υπερδοσολογία SNs, η μέγιστη ανεκτή δόση στους ανθρώπους είναι άγνωστη [82]. Σε εργαστηριακά πειράματα, οξείες από του στόματος υπερδοσολογία SNs σε ζώα (LD50) ήταν ως εξής: • σε αρουραίους 10.000 mg kg−1 (SSZ), • σε κουνέλια 2000 mg kg−1 (SSZ), • σε ποντικούς 5700 mg kg−1 (SSZ), 16.500 mg kg−1 3700–4200 mg kg−1 (SAD), 4500 mg kg−1 (STZ) [83].

(SCT),

Examples of side effects associated with SNs overdose in humans include nausea and skin hypersensitivity reactions. Other side effects, eg stomatitis, hemolysis, methemoglobinemia, hepatotoxicity and renal toxicity occur rarely. SN can cause interaction with other drugs, for example with methotrexate,

σουλφονυλουρίες, βαφαρίνη, μερκαπτοπουρίνη, κυκλοσπορίνη ή διδανοσίνη [84]. Κατά τη γνώμη μας, οι άμεσες τοξικές επιδράσεις που προκαλούνται από SN που εμφανίζονται στο περιβάλλον δεν φαίνεται να αποτελούν σημαντική απειλή για τη δημόσια υγεία. Πιθανές πιθανές περιπτώσεις άμεσων τοξικών επιδράσεων των SNs στον άνθρωπο μπορεί να είναι σποραδικές. Από την άλλη πλευρά, η εμφάνιση υπολειμμάτων SNs στα τρόφιμα, ιδιαίτερα σε περίπτωση παράνομης ή ακατάλληλης χρήσης αυτών των φαρμάκων, μπορεί να είναι πιο σοβαρό πρόβλημα. Σύμφωνα με τους Dolliver et al., τα υπολείμματα SNs στα τρόφιμα δεν αποτελούν απειλή ή/και δυσμενή επίδραση στην ανθρώπινη υγεία αλλά «ανάπτυξη και εξάπλωση της αντοχής στα αντιβιοτικά, η οποία είναι ένα σημαντικό πρόβλημα παγκοσμίως» [69]. 8. Αποικοδόμηση των SNs στους οργανισμούς και στο περιβάλλον Πιθανά προϊόντα του βιομετασχηματισμού και της αποδόμησης των SNs φαίνονται στο Σχ. 6. Μια λεπτομερής συζήτηση αυτών των διεργασιών παρουσιάζεται στις επόμενες ενότητες. 8.1. Μεταβολισμός SNs στα θηλαστικά Ένα μεγάλο μέρος της δόσης SNs απεκκρίνεται από τους οργανισμούς ως αμετάβλητες ενώσεις. Για παράδειγμα, το 75% του SMR θα μπορούσε να απεκκριθεί από το σώμα στη μητρική του μορφή [1]. Ωστόσο, γενικά, πάνω από το 80% μιας δόσης SN υφίσταται βιομετατροπή στα θηλαστικά. Ο βαθμός μετασχηματισμού κάθε ΣΝ εξαρτάται τόσο από τον τύπο του όσο και από τα χαρακτηριστικά του οργανισμού. Ο βιομετασχηματισμός των SN βασίζεται κυρίως στην οξείδωση, την ακετυλίωση ή την υδροξυλίωση στο άτομο αζώτου N4 ή τη γλυκουρονίωση των ατόμων αζώτου N1 - ή N4 - [1,10,11]. Υποτίθεται ότι, μετά την από του στόματος χορήγηση, το 50-70% της δόσης απεκκρίνεται στα ούρα ως N4-AcSNs και το 15-20% ως γλυκουρονίδια N1 [1,10]. Οι μεταβολίτες των SNs δεν έχουν υψηλή βιολογική δράση ως αμετάβλητα SNs. Ωστόσο, αυτή η δραστηριότητα θα μπορούσε εύκολα να αποκατασταθεί σε συνθήκες in vitro [11,85]. Οι συγκεντρώσεις μεταβολιτών διαφορετικών από αυτούς που αναφέρονται παραπάνω είναι μικρές και πιθανόν να μην είναι σημαντικές για το περιβάλλον. Ανασκοπήσεις πιθανών οδών βιομετατροπής SNs περιγράφηκαν στις εργασίες των Sukul and Spiteller [10] και García-Galán et al. [11]. 8.2. Βιοαποδομησιμότητα των SNs Οι απόψεις των ερευνητών σχετικά με τη βιοαποδομησιμότητα των SNs έχουν διχαστεί [1,4,10,17,24,86]. Η αιτία αυτού μπορεί να είναι οι διαφορές στη μικροβιακή δραστηριότητα της μήτρας, του εμβολίου που χρησιμοποιείται και των εφαρμοζόμενων μεθόδων που χρησιμοποιούνται για την αξιολόγηση της αποικοδόμησης του SN (Πίνακας 3). Η σταθερότητα των διαφόρων SN είναι επίσης διαφορετική. για παράδειγμα, το SDM είναι πιο (10x) ανθεκτικό στη βιοαποικοδόμηση από το STZ. Τα αποτελέσματα τυποποιημένων δοκιμών, όπως το ISO 11734:1995 και το OECD 301D, και η αξιολόγηση της μικροβιακής δραστηριότητας του εδάφους υποδηλώνουν ότι τα περισσότερα από τα SN δεν υφίστανται φυσική βιοαποδόμηση. Ένα από τα SN που περιγράφονται πιο συχνά στη βιβλιογραφία είναι το SMX, το οποίο έχει θεωρηθεί ως μη βιοαποδομήσιμη ένωση (σε καθαρό νερό, θαλασσινό νερό, φυσικό νερό και λύματα ή ενεργή λάσπη) σε 9 από 24 άρθρα [1,4,24, 86,89,91,97–99]. Σύμφωνα με τους Weifen et al. [114], παρουσία γαρίδας (Penaeus chinensis), η τιμή DT50 για το SMX είναι 5,68 ώρες. Οι Ingerslev και Halling-Sørensen [92] βρήκαν ότι, παρουσία μικροοργανισμών στην ενεργοποιημένη ιλύ, το DT50 των SNs είναι μόνο ~7 ώρες. Οι De Liguoro et al. [58] δήλωσε ότι, στην περίπτωση του SDT, το DT50 για μικροβιακή αποικοδόμηση σε φρέσκο στρώμα είναι ~ 1 ημέρα. Ομοίως, εξίσου ταχεία αποικοδόμηση της SDT έχει περιγραφεί από τους Wang et al. [87]. Αυτοί οι συγγραφείς έχουν παρατηρήσει μια αύξηση στην τιμή DT50 με αυξανόμενες αρχικές συγκεντρώσεις SDT σε νωπή και στείρα κοπριά. Σε αυτές τις περιπτώσεις, τα περισσότερα από τα SNs ενσωματώθηκαν σε μικροοργανισμούς και/ή υπέστησαν μόνο αναστρέψιμους μετασχηματισμούς, όπως η ακετυλίωση [11,85]. Η ταχεία εξαφάνιση των SNs στο έδαφος και την κοπριά θα μπορούσε να είναι αποτέλεσμα της δέσμευσης μεταξύ SNs και οργανικών ή ορυκτών σωματιδίων [85,88,93] ή μπορεί να προκληθεί από

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100.000

L(E)C50 (mg l-1) ή (mg kg -1)

10000 1000 100 10 1 0,1 0,01

Bacteria

Kiselalger

Tang

Aqua Gnitic Daria plante

Crustacea

Rotifer

You were fried

Fisk

Children (PNEC)

D. rerio (96 timer)

O. latipes (96 timer)

O. mykiss (EA)

T. platyurus (24 times) B. calciflorus (24 times) B. calciflorus (48 times) M. macrocopa (48 times)

C. dubia (7d)

C. doubt (48 hours)

H. attenuata (96 timers) D. walk (24 timers) D. walk (48 timers) D. walk (48 timers)

L. gibba (7δ)

Activate slam (20 times) AMES-test (72 times) C. meneghiniana S. capricornutum P. subcapitata (96 times) C. vulgaris (48 times) S. subspicatus (72 times) S. leopolensis (96 times)

C. freudii (24 times)

P. agglomerans P. aeruginosa (48 times) V. fischeri (5 min) V. fischeri (15 min) V. fischeri (30 min)

0,001

mama mals

Fig. 6. Comparison of SMX toxicity in selected test organisms.

photochemical processes (on the soil surface, in the presence of Fe and nitrate compounds) [107,115]. Most researchers have identified SN as poorly or non-biodegradable compounds in the environment (in clean water, surface water and soil with DT50 > 30 days) [24,74]. The fact that SNs appeared so frequently in the test samples can also be considered as evidence of their persistence in the environment. In our unpublished study on the biodegradation of SAD, STZ, SMX and SDZ applied to natural matrices, we found that in individual cases (STZ under aerobic conditions, in wastewater and wetlands) the DT50 was less than 2.5 days. In other cases, the DT50 values for SDZ, SAD and SMX were >5, >8 and 31 days, respectively. In our opinion, the reading stability data related to SN residues in the environment (especially for SMX) are generally much higher than the data reported by other researchers in the cited papers. The high frequency of SNs in the environmental samples described above can confirm this hypothesis. 8.3. Physico-chemical methods of SN decomposition The efficiency of SN decomposition using the most commonly used chemical and physico-chemical methods is shown in table 3. High efficiency of SN decomposition in wastewater was achieved using various advanced oxidation processes (AOP) [1, 4,161,101,1,1 4,101,101,101,1,1,4,1,100,000], such as the use of O3, Cl2 and ClO2 [1,101,116-118], the Fenton reaction [105,116] or photocatalytic processes [85,105,116]. Unfortunately, the implementation of these methods is expensive and can be harmful to the environment due to the formation of highly toxic intermediates [118]. In addition, a decrease in AOP efficiency was observed with increasing total wastewater contamination [108]. This fact made it difficult to directly apply these methods for the removal of SNs from manure. Table 3 shows examples of other methods used to remove pollutants from the aquatic environment without their degradation or transformation (non-destructive methods). SNs can be removed from wastewater with almost 100% efficiency by reverse osmosis [1,4,24,109,110]. However, with this method there may be a problem with wastewater containing concentrated solutions

toxins (including SN) [110]. In the case of substances that are resistant to biodegradation, there can be a local, dangerous increase in the concentration of these toxins in a small area [119]. Physico-chemical methods (especially AOP) can be effective and very useful in SN degradation. In our opinion, it is possible that their degradation in the environment is the result of photochemical reactions initiated by sunlight in the presence of natural photosensitizers, and not only the biodegradation process. 9. Removal of SN from wastewater There are different opinions about the efficiency of SN removal in conventional biological-mechanical treatment plants. Similar differences are visible in the assessment of the biodegradability of drugs (table 3). Onesios et al. [120] analyzed 49 cases of removal of SDZ, SDM, SMX and SPY from wastewater at a wastewater treatment plant (WWTP). Based on the analysis of recent publications, -280 to 100% of SNs were removed using activated sludge (AS) (Table 4). The average degree of SN removal in these cases was ~24%. According to data published in 2010, SMX was removed from selected wastewater treatment plants in Spain by 30–92% [35]. However, there are also cases where the SN concentration in the outflow was higher than in the inflow [4,86,134,138]. This phenomenon has been described in a pilot plant for wastewater treatment in Austria [90] and Switzerland [138]. This effect is probably caused by the hydrolysis of N4-AcSN present in the wastewater into parent SN [11]. A conclusion on this problem can be found in the data from the study of Turkdogan and Yetilmezsoy [109]. These authors estimated that 80% of used antibiotics enter the environment despite the use of different processes in sewage plants (based on data from Turkey, without taking SN into account). Importantly, a large fraction of SN can be adsorbed by biomass in wastewater plants [132] and can be returned to the environment. 10. Environmental risk assessment Most researchers used the method recommended by the European Medicines Evaluation Agency (EMEA) for environmental risk assessment. This method uses the results

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Table 3 Methods of biotransformation, degradation and other methods for SN removal. Matrix

Method

Biotransformation Human livestock manure

efficiency

DT50

99.0%) was from Aldrich and used as received. Hexane, sodium chloride and hydrochloric acid were of analytical grade. Deionized water (>18.0 m) was used for solution preparation and dilution. Commercial HA powder was obtained from Lemandou Chemicals Co., Ltd., China, which was derived from lignite (Sinkiang, China). Before use, HA was further purified according to the procedures recommended by the International Humic Substances Society (IHSS) [14]. The elemental composition of purified HA was analyzed on a Vario Micro elemental analyzer (Elementar, Germany). HA includes 48.3% C, 2.6% H, 29.3% O, 1.1% N, and 0.3% S. Kaolin (chemical grade) was obtained from Shanghai Qingpu Chemical CO. Ltd. was selected as modeling clay. The share of organic kaolin was 0.12%, and the cation exchange capacity (CEC) measured by the BaCl2-H2SO4 method (ISO 11260-1997) was 16.9 cmol kg−1. The BET specific surface area measured with a surface apparatus (Micromeritics Tristar 3000) was 19.4 m2 g−1. X-ray fluorescence analysis (Genesis, EDAX Inc.) shows that its main mineral elements were Si (59.9 wt%) and Al (36.7%). Other physicochemical properties associated with kaolin are listed in Table S1 in the Supplementary Materials.

2.2. Effect of HA-TX100 interaction on the sorption of HA and TX100 on kaolin. All sorption experiments were performed in triplicate using the batch-equilibrium technique in glass vials sealed with Teflon screw caps. Data processing and fitting (including isothermal model fitting) was performed using Origin v. 8.0. Initially, the sorption of HA and TX100 in the kaolin/water system was investigated. A total of 0.5 g of kaolin was mixed with 10 mL of HA (0-2500 mg L-1) or TX100 (0-10 mmol L-1) solution. The vials were vortexed on an end-over-end reciprocator for 48 h (25 ± 1 ◦ C) and centrifuged at 4000 revolutions per minute for 10 minutes. That

The supernatant was filtered through a 0.45 µm cellulose acetate membrane and subjected to HA or TX100 analysis. The effect of HA-TX100 interaction on TX100 and HA adsorption on kaolin was further investigated in two different ways (0.5 g kaolin in 10 mL solution): with a fixed total HA concentration of 400 mg L-1 and varying total TX100 concentration from 0 – 10 mmol L−1, or with a fixed total TX100 concentration of 2 mmol L−1 and different total HA concentrations of 0–2500 mg L−1. The pH of the sludge was adjusted to 7.0 ± 0.2 with appropriate concentrations of HCl or NaOH (generally 1 mol L-1, while 10 mol L-1 HCl or NaOH was used for samples with high HA concentrations). In addition, 0.01 mol L-1 NaCl was included as background electrolyte. Equilibrium concentrations of HA and TX100 in water were analyzed for each sample. 2.3. Effect of HA or TX100 on HCB sorption on kaolin In this section, the effect of HA addition on HCB sorption on kaolin was investigated for the first time. A total of 0.5 g of kaolin was mixed with 10 ml of HA solution (0-2500 mg L-1, final pH 7). Then, 40 µL of the acetone-HCB solution was added to the vials with a microsyringe before shaking (acetone fraction was below 0.5%). The total HCB concentration was 0.5 mg L-1 (or 10 mg kg-1 kaolin). After equilibration and centrifugation, the supernatant was filtered through a 0.45 µm cellulose acetate membrane. The HCB in the supernatant was then extracted with hexane by liquid-liquid extraction in 11 mL vials (with extraction ratio 1:2, shaken on a shaker for 2 hours) and analyzed by gas chromatography (GC). Meanwhile, HA in the filtrate was measured. Second, the effect of TX100 on HCB adsorption was investigated according to the same procedure as described above. The concentration of TX100 used was 0-10 mmol L-1. Equilibrium concentrations of TX100 and HCB were determined for each sample. 2.4. Separation of HCB in the HA/TX100/water solid system Predetermined amounts of HA and TX100 stock solutions (HA 5 g L-1 and TX100 20 mmol L-1) were pipetted into 50 mL glass bottles and diluted with deionized water to 20 mL . The amounts of HA added were 0-2.5 g L-1 and the total concentration of TX100 was 2 mmol L-1. The pH of the mixture was adjusted to 2-3 with HCl to completely precipitate the HA. Then 80 µL of acetone-HCB solution was added before mixing. The total concentration of HCB was 0.5 mg L−1. After equilibration and centrifugation, the supernatant was filtered through a 0.45 µm cellulose acetate membrane and subjected to TX100 and HCB analysis. 2.5. Separation of HCB in the HACK/TX100/water system. The sorption of HCB onto HACK was additionally investigated (detailed preparation procedures are provided in Supplementary Materials) in the presence of TX100. A total of 0.5 g of HACK with different amounts of HA coating (0.25-5%) was mixed with 10 ml of TX100 solution (0-10 mmol L-1). The pH of the slurry was then adjusted to 7.0 ± 0.2 or 3.0 ± 0.2 (pH 3.0 may avoid the dissolution of HA by HACK, thus better revealing the role of solid HA in the degradation of TX100 and HCB). Then, 40 µL of the acetone-HCB solution was added to the vials with a microsyringe before shaking (the final acetone fraction was below 0.5%). Equilibrium concentrations of TX100 and HCB were determined for each sample. 2.6. Chemical analysis Aqueous TX100 was analyzed by high performance liquid chromatography (Hitachi L7100, Japan) equipped with

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Fig. 1. (a) TX100 sorption isotherm at 0 and 400 mg L−1 HA. (b) influence of the amount of HA on the sorption of TX100 (2 mmol L−1) on kaolin.

L-7420 ultraviolet-visible (UV-vis) detector and Agilent ZORBAX Eclipse XDB-C18 column (Agilent, USA). The wavelength is set to 223 nm. The mobile phase was 90% methanol plus 10% water at a flow rate of 1.0 mL min-1. HCB in hexane was determined on a Hewlett-Packard 6890 GC equipped with an electron capture detector and a ZB-5 capillary column (Phenomenex, USA). Detailed information about the GC procedure is included in our previous study [15]. The concentration of HA in water was measured with a UV-vis spectrophotometer Cary 50 (Varian, USA) at 254 nm (although TOC analysis is often used to quantify humic substances (including HA), here the co-presence of TX100 especially above 2 mmol L−1 can seriously affect to TOC analysis of HA, but the much higher absorbance of HA under UV-254 than TX100 suggests that spectrophotometer measurement may be more appropriate). The linear range of the HA operating curve was 0-24 mg L-1 (r = 0.9999). HA absorption in the absence of TX100 can be obtained directly. Based on the preliminary observation that the absorbance of the mixture of TX100 and HA at 254 nm is additive, the HA content can be deduced by subtracting the absorbance contributed by TX100 from the absorbance of the TX100-HA mixture (absorbance of TX100 can be calculated from the operating curve at 254 nm and the corresponding concentration is obtained using HPLC). 3. Results and discussion 3.1. Effects of SOM on TX100 sorption in kaolin-water system. Fig. Fig. 1a shows the sorption isotherms of TX100 in the absence and presence of HA (400 mg L-1) in the kaolin/water system. The two isotherms can be well adjusted by the Langmuir model in the range 0-10 mmol L-1, which corresponds to those published in the literature [7,16]. A comparison of the maximum sorption amounts of TX100 between HA and HA systems (24.0 vs. 19.1 mmol kg−1) shows that

the presence of HA at 400 mg L-1 increased the absorption of TX100. Several researchers have reported increased sorption of surfactants into soil as a result of SOM [17-19]. Zhu et al. [20] found that the sorption capacity of HA for TX100 was almost an order of magnitude higher than that of kaolin. Zhang et al. [21] also reported that HA showed much higher adsorption capacity and partition coefficient (Kd) of TX100 than native soil. Furthermore, as shown in Fig. 1b, for the HA concentration range observed, TX100 sorption first increased and then decreased with a maximum HA concentration of about 400 mg L-1. The initial increase in TX100 sorption was highly correlated with the surfactant distribution in the clay-bound HA. However, as added HA increased above 400 mg L-1, the amount and more importantly, the proportion of dissolved HA increased significantly (as indicated by the decrease in Kd values in Table S2). Dissolved HA can compete with clay-bound HA for TX100 compartmentalization, since the HA molecule contains a hydrophilic and hydrophobic structure that is very similar to surfactants. Another explanation could be that dissolved HA can reduce the adsorption of TX100 by forming HA-TX100 complexes, similar to the complexation of HOC in DOM [10,22,23]. Lee et al. [22] found that the TX100 sorption isotherm in a Florida peatland had a skewed Gaussian shape with a sharp decrease in the TX100 sorption coefficient at an equilibrium concentration above 1000 mg L-1. It was suggested that at higher TX100 concentrations, more SOM was dissolved and DOM was expected to enhance TX100 “dissolution” and reduce the probability of TX100 sorption [22]. Inspection of Fig. 1b further reveals that even at relatively high concentrations of dissolved HA (100-400 mg L-1), the sorption of TX100 was greater than the sorption at HA = 0, indicating significant binding of TX100 to binding with HA. kaolin. It can also be estimated from Fig. 1b that the critical amount of HA when dissolved HA exceeded the clay-bound fraction for TX100 separation was about 1500 mg L-1. As a result, the effect of SOM on nonionic surfactant sorption is expected to depend on the SOM content, and only at a content high enough to induce a significant amount of DOM can SOM reduce surfactant sorption. they have a positive effect in SER.

3.2. Effects of TX100 on HA adsorption in kaolin/water system. The distribution of HA in the kaolin/water system is shown in Fig. 2a. The HA sorption isotherm in the system without TX100 can be described as a two-phase linear relationship: with a sharp increase for dissolved HA concentration of 0-7 mg L-1 and a subsequent slowing increase in the range 7-360 mg L-1 −1. Furthermore, linear regression could be applied to HA sorption in the presence of TX100 when the HA concentration in water varied from 29 to 610 mg L-1. Examining Fig. 2a reveals that the co-presence of TX100 at 2 mmol L-1 reduces the adsorption of HA on kaolin. Furthermore, the slope of the HA isotherm in the presence of TX100 is lower than that of HA in the absence of TX100 (5.7 vs. 7.7), indicating an increased decrease in HA sorption as dissolved HA increases. Furthermore, Fig. 2b shows a uniform decrease in HA sorption with increasing TX100 concentration (0-8 mmol L-1). In particular, the presence of 8 mmol L-1 TX100 could reduce HA uptake by 38% compared to the system without TX100. Similar observations were also reported by Cheng and Wong [13], where the presence of Tween 80 at 150 mg L-1 dramatically reduced soil DOM uptake. Preferential binding of Tween 80 molecules in soil compared to DOM is assumed to be relevant to the above results [13]. Despite the lower calculated partition coefficients for TX100 (2–34 L kg−1 at 1–10 mmol L−1, Table S1) compared to HA (23.5–80 L kg−1 at 350–1000 mg L− 1, Table S1 ) ), it might still be reasonable to suggest that it is competitive

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Fig. 2. (a) HA sorption isotherm at 0 and 2 mmol L−1 TX100. (b) influence of TX100 concentration on the sorption of HA (400 mg L-1) on kaolin.

Adsorption between the two molecules on the clay surface mainly contributed to the decreased absorption of HA as aqueous TX100 increased. 3.3. Partitioning of HCB in the HACK/TX100/water system Although the effect of TX100-HA interaction on the adsorption of TX100 and HA on kaolin has been confirmed, the resulting effects on the partitioning of HCB in the clay/water system remain unclear. In fact, in the complex surfactant/SOM/water/clay system, a number of sorption and complexation processes can occur that can affect the distribution behavior of HCB. To be more specific, these interactions include (1) the formation of a DOM-HCB complex that can inhibit HCB sorption [10,24]. (2) the role of clay-bound SOM as a hydrophobic sorption site for HCB [23,25,26]. (3) the distribution of HCB micelles in TX100, which can effectively reduce the sorption capacity of HCB [3,27]. (4) the effect of TX100-HA interaction on the adsorption of TX100 and HA on the clay surface, which may further affect the distribution of HCB. (5) possible sorption of HCB by clay bound TX100 and mineral [27-29]. Considering the complexity of HOC distribution in the simultaneous presence of TX100 and SOM, the sorption of HCB on individual SOM and TX100 systems was first investigated. In Figure 3a, it can be seen that the adsorption of HCB was negatively correlated with the amount of added HA at HA > 250 mg L-1. The fraction of adsorbed HCB steadily decreased from 1.0 to 0.5 as HA increased to 2500 mg L-1. With increasing added HA, dissolved HA became more abundant, inhibiting HCB absorption. However, in the low HA range of 0-150 mg L-1, HCB adsorption increased slightly. It is generally accepted that HA could bind HOC through hydrophobic interactions into HA-HOC complexes [10,24,30], with soil-bound HA adsorbing HOC from solution, while dissolved HA desorbs HOC from soils. On the other hand, the formation of a micellar structure of HA due to the aggregation of molecules is also proposed as a mechanism of HA solubilization/enhanced desorption for HOC [9,27]. Furthermore, Fig. 3b

Fig. 3. Effect of (a) HA and (b) TX100 on HCB sorption onto kaolin separately. HCB 0.5 mg L-1, pH 7.0.

that HCB sorption showed a constant decrease with increasing added TX100 (0-2 mmol L-1, Figure 3b). In particular, when 2 mmol L-1 TX100 was added, the proportion of adsorbed HCB was ca. 0.4. It is suggested that only at a total concentration of surfactant above the critical desorption concentration (CDC), i.e. the concentration of surfactant in water above the CMC, a reliable improvement of HCB desorption with surfactants can be achieved [3,31]. The CDC for TX100 here can be estimated from the following equation: CDC = CMC + CsCMC where CsCMC is the soil concentration of TX100 when the corresponding water concentration is at CMC, which can be calculated from CsCMC = 19.1 CMC/( 0.085 + CMC ) ( figure 1a). By substituting the measured TX100 CMC value of 0.1 mmol L-1 into the above equations, the estimated CDC is 0.62 mmol L-1. Consequently, no obvious HCB desorption was noted at a total TX100 concentration of 0.5 mmol L-1, while significant HCB desorption was achieved at a TX100 dose above 1 mmol L-1. The distribution of HCB in the HACK/TX100/water system and the effects of HA-bound clay and dissolved HA on the adsorption of HCB onto HACK are illustrated in the figure. 4a-c. From fig. In Figure 4a, it is clear that the sorption of TX100 and HCB on solid HA was positively correlated with HA content. As the HA concentration increased from 0 to 2.5 g L-1, aqueous TX100 (initially 1 mmol L-1) decreased significantly to almost unmeasurable, and accordingly, HCB sorption increased steadily (initial aqueous concentration was 0.5 mg L-1). to 1.0. The reduced aqueous concentration of TX100 was directly related to the high affinity of HA for TX100, as the Kd values for TX100 here were calculated as (0.6-8) × 103 L kg-1, which is 1.8-2.9 orders of magnitude higher of TX100 in kaolin (10.9 L kg−1 in TX100 2 mmol L−1). However, the reduced concentration of aqueous HCB was mainly related to the strong affinity of HA and TX100 bound to HA [27]. The binding constant (Kb) of the HA-HCB complex was calculated as 5.5 × 104 L kg-1 using the solubility enhancement method [24,32]

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Fig. 5. Evaluation of factors that contributed to HCB sorption, where "HA only" and "TX100" represent only HCB sorption as an intrinsic effect of HA (Figure 3a) and 2 mmol L-1 TX100 (Figure 3b), respectively , "Theoretical effect of TX100 in the system TX100/HACK" represents the theoretical sorption of HCB TX100 under the influence of the HA coating, calculated from the water content of TX100 (Figure 4c) and the intrinsic effect of TX100 in HCB-adsorption Figure 3).

the amount of HA coating 0.5–5% (Fig. S2, the slope of the linear “SOM content” was less than 0.5). As a result, the formation of the HA-HCB aqueous complex became more intense. In addition, the reduced absorption of TX100 in the presence of HA may be further responsible for the reduction of HCB adsorption. As indicated in Fig. 4c, the dissolved fraction of TX100 increased from 1.3 to more than 1.6 mmol L-1 as the amount of HA coating increased, which is slightly different from the results in Fig. 1b. A likely explanation for this difference was that for HACK the clay surface was pre-occupied by HA and that the exchange of TX100 for bound HA was somewhat more difficult than competitive sorption with dissolved HA (Section 3.1). Finally, a possible synergistic effect of the HA-TX100 complex on HCB solubilization and reduced adsorption in HACK/surfactant/water systems cannot be excluded [13]. 3.4. Contribution of SOM and TX100 to HCB sorption

Fig. 4. Distribution of TX100 and HCB in (a) solid HA/water, (b) HACK/water (pH 3.0) and (c) HACK/water (pH 7.0) system. The initial concentrations for HCB and TX100 were 0.5 mg L-1 and 2 mmol L-1, respectively.

(more detailed in the notice of Fig. S1 in Supplementary Materials), suggesting a rather strong interaction between HA and HCB. In addition, the reduced aqueous concentration of TX100 as a function of solid HA could also contribute to the increased sorption of HCB given the poor inherent solubility of HCB in water. Similar trends for TX100 and HCB sorption were also observed in the HACK/water system at pH 3 (Figure 4b). Note that much lower aqueous TX100 (about 0.9 mmol L-1) and higher HCB adsorption (above 0.6) were recorded at a very low HA content of 0.25% (Fig. 4b), which mainly attributable to the absorption of kaolin. The important role of inorganic components in the distribution of surfactants or HOCs, especially at low organic carbon content, has been considered by other researchers [33,34]. Fig. Fig. 4c illustrates the distribution of TX100 and HCB at pH 7.0 in the HACK/water system. In contrast to the trends at pH 3.0 (Figure 4b), the fraction of adsorbed HCB decreased from 0.47 to 0.26 with HA coverage varying from 0.25% to 5%. The decrease in HCB absorbance was consistent with the above observations in the HA/kaolin system (Figure 3a). It is worth noting that although increasing the HA coating may lead to an immediate increase in the SOM content of kaolin, the proportion of dissolved HA is clearly greater than the proportion bound to kaolin in the whole area

Based on the above results of the distribution behavior of HCB in individual systems, in this section we will further summarize and compare the individual and combined roles of HA and TX100 in HCB sorption. For this purpose, Figure 5 and Table 1 were constructed. The dotted curve (marked as "only HA", i.e. Figure 3a) and the dashed curve (marked as "only TX100", i.e. Figure 3b) represent the adsorption of HCB on clay, when HA (125-2500 mg L-1) and TX100 (2 mmol L-1) were added individually. The solid curve (labeled “HA and TX100 both”, i.e. Figure 4c) shows the adsorption of HCB due to the simultaneous presence of TX100 and HA. The dashed curve (labeled "Theoretical effect of TX100 on the TX100/HACK system") is the theoretical HCB adsorption induced by 2 mmol L-1 TX100 on the TX100/HACK system, where the effect of added aqueous TX100 (as a result of and ) from the HA coating , section 3.3 and Fig. 4c) was involved in the sorption of HCB (see the detailed removal procedure from the notes in Table 1). It should be noted that a neutral pH of 7.0 has been set for all the mentioned systems, so the influence of pH can be ignored. Examining Fig. 5 shows that TX100 had a dominant role in reducing HCB adsorption. In addition to increasing the solubility of HCB from native aqueous TX100 (in the HA-free system), the reduced TX100 sorption due to the concomitant presence of HA may further reduce HCB sorption. This additional fraction of adsorbed HCB was expected to be about 0.04, as revealed by comparing the "TX100 observed" and "TX100 theoretical in the TX100/HACK system" results (Table 1). However, the effect of HA coating on HCB separation was doubly dependent on the amount of coating. Although there was a dramatic decrease in HCB absorption as a function of increased HA coverage across the range, the presence of HA was found to promote HCB adsorption

Οι J. One et al. / Journal of Hazardous Materials 196 (2011) 79–85

Table 1. Contribution of TX100 and HA to HCB desorption from HACK. HA occupancy (%)

0,5 1,0 2,0 3,0 4,0 5,0

The fraction of HCB adsorbed on different systems, observed Ce/C0 HA

TX100 observed

TX100 theoretically on the HACK/TX100b system

HA & TX100 both - theoretically Mr

Sum-HA & TX100 separate

0,03 0,24 0,26 0,40 0,44 0,50

0,59 0,59 0,59 0,59 0,59 0,59

0,61 0,63 0,63 0,63 0,63 0,63

0,53 0,54 0,58 0,62 0,65 0,74

0,61 0,83 0,84 0,99 1,03 1,10

Observed desorption fraction of HCB from HA alone (results in Figure 3a). Similarly for "TX100-observed" (results in Figure 3b). The theoretical fraction of HCB desorption as a result of 2 mmol L-1 TX100 in the TX100/HACK system, calculated by substituting the value of aqueous TX100 (Fig. 3c) into the following equation relating HCB adsorption to the concentration of aqueous TX100: si

FHCB =

0,109 + 1,08 1 + 1,95 CeTX100 0,946

where FHCB is the fraction of HCB adsorbed on HACK, CeTX100 is the aqueous concentration of TX100. c Observed HCB desorption fraction as an effect of HA coating and TX100 addition. d Theoretical fraction of HCB desorption from TX100 and HA, i.e. sum of HCB desorption fraction from HA and TX100 individually.

at smaller amounts of coating, if you take the "TX100 only" result as a reference (fig. 5). As discussed above, the competition between aqueous and clay bound HA determines the effect of the HA coating on HCB adsorption [10,23], independent of TX100 interference. Note that the intersection of the “HA & TX100 both” curve and the “TX100 only” curve (P1 in Figure 5 ) represents the balance between the facilitating and hindering roles of the HA coating on HCB adsorption. It was hypothesized that only at HA coverage greater than 2.4% dissolved HA outperformed clay-bound HA for HCB separation. Excessive desorption of HCB from clay bound HA could be achieved due to this dissolved HA compared to desorption of HCB by TX100 alone. If we really take into account the additional effect of TX100 adsorption reduction, the threshold value will shift towards higher HA coverage, as shown in Figure 5 (P2), which means that positive effects in reducing HCB HA adsorption are expected at coating amounts greater than 3 .3%. Inspection of Fig. 5 further reveals that P1 is indeed the critical point where the combined effects of HA and TX100 on HCB adsorption are equal to the effects of TX100 alone. This means that at HA coverage greater than P1, the simultaneous presence of TX100 and HA resulted in less HCB sorption than either HA or TX100 alone. However, as indicated in Table 1, the combined effect on reducing HCB adsorption was less than the sum of TX100 and HA alone. The results apparently conflict with previous observations that a synergistic effect on HOC desorption from soil was reported when Tween 80 and DOM were added together [13]. However, note that the soil texture (loamy sandy soil) and agent addition procedures (simultaneous addition of Tween 80 and DOM system) in it would likely cause less DOM sorption compared to HACK in our study. Since the fractions of HCB "desorption" were 60% and above 40% for TX100 (2 mmol L-1) and HA (at ≥ 2% HA coverage) alone (Table 1), no synergistic or even an additive effect from the combination of TX100 and HA. while, as reported by Cheng and Wong [13], even when a synergistic effect was achieved, the maximum desorption ratio was only 16.2% and 10.9% for phenanthrene and pyrene, respectively. We hypothesized that a similar synergistic or additive effect of the simultaneous presence of TX100 and HA can be achieved if an appropriate concentration of TX100 and HA is applied (at least the sum of HCB desorption efficiency with TX100 and HA alone is clearly below 100%). .

4. Conclusions A deeper understanding of the combined effect of surfactants, SOM and DOM on HOC sorption in the soil/water system is needed

better predict the effectiveness of SER. Here, we investigated the distribution of HCB in the HACK/water system in the presence of TX100. The main conclusions can be summarized as follows: (1) A remarkable effect of TX100-HA interaction on the adsorption of TX100 and HA on clay was observed. It was found that the addition of HA in a lower and higher amount than 1500 mg L-1 increases and decreases the absorption of TX100 on kaolin, respectively. Furthermore, the presence of TX100 suppressed the sorption of HA onto kaolin in the entire observed concentration range of 0.5-2 mmol L-1. The presented results suggest that for soils with a high organic matter content, the interaction of surfactant and SOM can be beneficial for SER. Additionally, a suitably high dosage of surfactant may be preferable to solubilize SOM and effect such further desorption. (2) TX100 mainly contributed to the desorption of HCB in the HACK/TX100/water system. DOM also showed an encouraging improvement in HCB desorption. However, coated HA showed a negative and positive effect on TX100-enhanced HCB desorption at coating amounts below and above 2.4%, respectively. The combined effect of HA and TX100 on HCB desorption was less than the sum of TX100 and HA alone. (3) It should be noted that HACK may or may not fully represent the entire terrain in nature. However, the obtained results are still of practical importance for SER, since HA is often closely related to clay minerals in soil, and both HA and clay minerals are the primary soil compounds that determine the absorption of surfactants and HOCs, i.e. . (4) Our study suggests that for soils with a high organic matter content, a combined effect of SOM and surfactants on HOC removal can be expected, implying a higher efficiency of SER or a lower dosage of surfactant may be required. However, a critical SOM content was detected here that distinguishes the positive role of SOM from the negative in SER, and the value may vary from site to site depending on a range of factors such as soil characteristics, SOM, surfactants and EOE. In addition, pH as an important environmental variable can affect the amount of DOM and thus affect the SER process. However, in order to consider the simplification of the system, we only chose pH 7 as a typical circumstance in this study. As a result, further studies are still needed to focus on the combined effect of surfactants and SOM on SER for real soils with different organic matter contents, a wider range of pH conditions, and HOC variations.

Οι J. One et al. / Journal of Hazardous Materials 196 (2011) 79–85

Acknowledgments This work was supported by the National Natural Science Foundation of China (Grants 20777024), National High-Tech Research and Development Program (863) (2009AA063103), and Shanghai Tongji Gao Tingyao Environmental Protection Sci. & Tech. Development Fund. We thank the Analytical and Testing Center of Huazhong University of Science and Technology for assistance in characterizing kaolin and HA. Appendix A. Supplementary data Supplementary data related to this article can be found in the online version at doi:10.1016/j.jhazmat.2011.08.072. Literature [1] M. Svab, M. Kubala, M. Muellerova, R. Raschman, Soil ushing by Surfactant Solution: Pilot-scale demonstration of full technology, J. Hazard. Mater. 163 (2009) 410-417. [2] S.H. Yuan, Z. Shu, J.Z. Wan, X.H. Lu, Enhanced desorption of hexachlorobenzene from kaolin with single and mixed surfactants, J. Colloid Interface Sci. 314 (2007) 167-175. [3] K. Yang, L.Z. Zhu, B.S. Xing, Improved soil washing of phenanthrene with mixed solutions of TX100 and SDBS, Environ. Sci. Technol. 40 (2006) 4274-4280. [4] C.N. Mulligan, R.N. Yong, B.F. Gibbs, Surfactant-enhanced remediation of contaminated soils: a review, Eng. Geol. 60 (2001) 371-380. [5] W. Huang, W.J. Weber, A distributed reactivity model for soil and sediment sorption. 10. Relationship between desorption, hysteresis and chemical properties of organic domains, Environ. Sci. Technol. 31 (1997) 2562-2569. [6] R. Chefetz, A.P. Deshmukh, P.G. Hatcher, Sorption of pyrene by natural organic matter, Environ. Sci. Technol. 34 (2000) 2925-2930. [7] M.J. Salloum, M.J. Dudas, W.B. McGill, S.M. Murphy, Adsorption of surfactants on soils and geological samples of different mineralogical and chemical properties, Environ. Toxicol. Chem. 19 (2000) 2436-2442. [8] F.J. Ochoa-Loza, W.H. Noordman, D.B. Jannsen, M.L. Brusseau, R.M. Maier, Effect of clay, metal oxides and organic matter on the sorption of rhamnolipid biosurfactants from soil, Chemosphere 66 (2007) 1634-1642. [9] P. Conte, A. Agretto, R. Spaccini, A. Piccolo, Soil remediation: humic acids as natural surfactants in the washing of highly polluted soils, Environ. Pollution. 135 (2005) 515-522. [10] M. Rebhun, F. De Smedt, J. Rwetabula, Dissolved humic substances for cleanup of sites contaminated with organic pollutants. Predictions of binding-sorption models, Water Res. 30 (1996) 2027-2038. [11] H.H. Cho, J. Choi, M.N. Goltz, J.W. Park, Combined effect of natural organic matter and surfactants on the apparent solubility of polycyclic aromatic hydrocarbons, J. Environ. qual. 21 (2002) 275-280. [12] H. Lippold, U. Gottschalch, H. Kupsch, Joint influence of surfactants and humic matter on PAH solubility. Do mixed micelles occur? Chemosphere 70 (2008) 1979-1986. [13] K.Y. Cheng, J.W.C. Wong, Combined effect of the nonionic surfactant Tween 80 and DOM on the behavior of PAHs in the soil-water system, Chemosphere 62 (2006) 1907-1916.

[14] B. Wen, J.J. Zhang, S.Z. Zhang, X.Q. Shan, S.U. Khan, B.S. Xing, Sorption of phenanthrene to humic acid and humic fractions of soil, Environ. Sci. Technol. 41 (2007) 3165-3171. [15] J.Z. Wan, S.H. Yuan, K.T. Mak, J. Chen, T.R. Li, L. Lin, X.H. Lu, Enhanced leaching of HCB-contaminated soil with methyl-beta-cyclodextrin combined with ethanol, Chemosphere 75 (2009) 759-764. [16] P. Wang, A.A. Keller, Distribution of hydrophobic organic compounds in surface-active soil-water systems, Water Res. 42 (2008) 2093-2101. [17] S. Paria, Surfactant-enhanced remediation of organically contaminated soil and water, Adv. Colloid Interface Sci. 138 (2008) 24-58. [18] I.F. Patterson, B.Z. Chowdhry, P.J. Carey, S.A. Leharne, An investigation of the adsorption of triblemers of ethylene oxide and propylene oxide onto soil, J. Contam. Hydrol. 40 (1999) 37-51. [19] C.K.J. Yes, L.C. Lin, Sorption and desorption kinetics of surfactants TX-100 and DPC in different soil fractions, J. Environ. Sci. Harmful/dangerous. Supt. Surround. Meadow. 38 (2003) 1145-1157. [20] L.Z. Zhu, K. Yang, B.F. Lowe, B.H. Yuan, Multicomponent statistical analysis of the effect of sediment/soil composition on the adsorption of a nonionic surfactant (Triton X-100) to natural sediments/soil, Water Res. 37 (2003) 4792-4800. [21] G.Z. Zhang, H. Hu, W.L. Sun, J.R.J. Ni, Sorption of Triton X-100 on soil organic matter fractions: kinetics and isotherms, Environ. Sci. 21 (2009) 795-800. [22] J.F. Lee, R.M. Liao, C.C. Kuo, H.T. Yang, C.T. Chiou, Effect of a nonionic surfactant (Triton X-100) on the partitioning of contaminants between water and several soil solids, J. Colloid Interface Sci. 229 (2000) 445-452. [23] Y. Laor, W.J. Farmer, Y. Aochi, P.F. Strom, Binding and sorption of phenanthrene to dissolved and mineral-bound humic acid, Water Res. 32 (1998) 1923-1931. [24] C. Rav-Acha, M. Rebhun, Binding of organic solutes to dissolved humic substances and its effect on adsorption and transport in the aquatic environment, Water Res. 26 (1992) 1645-1654. [25] S. Kleineidam, H. Rügner, B. Ligouis, P. Grathwohl, Facies of organic material and equilibrium sorption of phenanthrene, Environ. Sci. Technol. 32 (1999) 1637-1644. [26] Y. Ran, K. Sun, Y. Yang, B.S. Xing, E. Zeng, Strong phenanthrene sorption of concentrated organic matter in soil and sediments, Environ. Sci. Technol. 41 (2007) 3952-3958. [27] S.O. Krava, M.A. Schlautman, E.R. Carraway, Distribution of hydrophobic organic compounds on adsorbed surfactants. 1. Experimental studies, Environment. Sci. Technol. 32 (1998) 2769-2775. [28] J. Hur, M.A. Schlautman, Effects of mineral surfaces on the distribution of pyrene in well-characterized humic substances, J. Environ. qual. 33 (2004) 1733-1742. [29] Y.J. Zhang, D.Q. Zhu, H.X. Yu, Sorption of aromatic compounds on mineral clay and modeling of humic substance-clay complexes: effects of solute structure and exchangeable cation, J. Environ. qual. 37 (2008) 817-823. [30] F. De Paolis, J. Kukkonen, Binding of organic pollutants to humic and fulvic acids: influence of pH and structure of humic material, Chemosphere 34 (1997) 1693-1704. [31] W. Zhou, L. Zhu, Effectiveness of surfactant-enhanced desorption for contaminated soil depending on the characteristics of the components of the soil-surfactant-PAH system, Environ. Pollution. 147 (2007) 66-73. [32] A.D. Site, Factors influencing the sorption of organic compounds on natural sorbent/water systems and sorption coefficients of selected pollutants, J. Phys. Chem. Ref. data 30 (2001) 187-439. [33] J.C. Joo, C.D. Shackelford, K.F. Reardon, Sorption of nonpolar neutral organic compounds on sand coated with humic acid: contribution of organic and mineral components, Chemosphere 70 (2008) 1290-1297. [34] X. Shi, L. Ji, D. Zhu, Investigation of the roles of organic and inorganic soil components in the sorption of polar and nonpolar aromatic compounds, Environ. Pollution. 158 (2010) 319-324.

Journal of Hazardous Materials 196 (2011) 86–92

Content is available on SciVerse ScienceDirect

Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Geopolymers Made from Plasma Treated Continuous Air Pollution Control (APC) Residues: Properties and Characterization of the Binder Phase Ioanna Kourti a, Amutha Rani Devaraj a,b, Ana Guerrero Bustos c, David Deegan d, Aldo R. Boccaccini b,e, Christopher R. Cheeseman a ,∗ a

Department of Civil and Environmental Engineering, Imperial College London, London SW7 2AZ, UK Department of Materials, Imperial College London, London SW7 2BP, UK c Institute of Construction Science Eduardo Torroja (CSIC), C/Serrrano Galvache, 4, 28033 Madrid, Spain d Tetronics Ltd., South Marston Business Park, Swindon, Wiltshire SN3 4DE, UK e Institute for Biomaterials, Department of Materials Science and Engineering, University of Erlangen-Nuremberg, Cauerstr. 6, 91058 Erlangen, Germany β

article

i n f o

Article history: Received February 14, 2011 Received in revised form August 30, 2011 Accepted August 31, 2011 Available online September 6, 2011 Keywords: Geopolymer Plasma combustion Energy Waste APC

a b s t r a c t Air pollution control (APC) residues are mixed with glass-forming additives and processed using DC plasma technology to produce high-calcium aluminosilicate glass (APC glass). This has been used to form geopolymer-glass composites that exhibit high strength and density, low porosity, low water absorption, low leaching, and high acid resistance. The composites have a microstructure consisting of residual unreacted APC glass particles embedded in the geopolymer composite and C–S–H binder gel phase and behave as particle-reinforced composites. The work shows that materials made from APC residues treated with DC plasma have the potential to be used to make high-quality prefabricated products. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Air pollution control systems in energy-from-waste (EfW) plants that burn municipal solid waste (MSW) produce air pollution control (APC) residuals. It is a hazardous waste with absolute registration in the European waste list (19 01 07*) and contains fly ash, excess lime, carbon, relatively high concentrations of volatile heavy metals and soluble salts, especially chlorides, which can be washed away. They also contain traces of organic substances, including dioxins and furans. DC plasma technology provides a sustainable treatment of APC residues that meets EU waste policy objectives, as it is a reuse/recovery option higher up the waste management hierarchy than alternatives [1,2]. In the DC plasma treatment process, APC residues are combined with glass-forming additives and melted to obtain APC-inert glass [2]. There is a growing interest in the development of sustainable building products that contain recycled materials. Recycling APC glass would have significant economic and environmental benefits

∗ Corresponding author. Phone: +44 207 594 5971; fax: +44 207 823 9401. E-mail address:[email protected](C.R. Cheeseman). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.08.081

would help make DC plasma treatment of APC debris commercially attractive. Previous work has investigated the use of APC glass in glass ceramics [3,4] and sintered ceramic tiles [5]. Geopolymers are synthetic aluminosilicates composed of tetrahedral silicon (SiO4) and aluminum oxide (AlO4), bonded by common oxygen atoms [6]. Their formation is based on the chemistry of alkali-activated inorganic binders and involves the chemical reaction of geopolymeric precursors, such as aluminosilicate oxides with polysilicate alkalis, to form polymeric Si-O-Al bonds [7]. The negative charge of Al3+ in fourfold coordination is balanced by the presence of positive ions such as Na+, K+ and Ca2+ in the skeletal cavity [8]. The empirical formula for geopolymers is therefore: Mn (–(SiO2 )z –AlO2 )n ,wH2O

(1)

where M is a cation such as Na+, K+ or Ca2+, z is 1, 2 or 3 and n is the degree of polycondensation. Growers are associated with low CO2 emissions compared to Portland cement [9]. Using APC waste glass to create geopolymers would allow for low-carbon recycling that does not involve heat treatment. The microstructure and properties of geopolymers are determined by the raw materials used and can have a high wound

I. Kurti et al. / Journal of Hazardous Materials 196 (2011) 86–92 Table 1 Chemical composition of APC glass. Oxide

Composition (% by weight)

Na2 O MgO Al2 O3 SiO2 P2 O5 K2 O CaO TiO2 Mn3 O4 Cr2 O3 Fe2 O3 Cl

2,88 2,31 14,78 41,10 0,77 0,03 32,59 1,19 0,23 0,06 4,07 2,5

compressive strength, low shrinkage, fast or slow hardening, good acid and fire resistance, and low thermal conductivity [10-12]. The use of APC glass in geopolymers results in geopolymer-glass composites, where the remaining APC glass particles act as rigid inclusions in the geopolymer matrix [13]. The presence of calcium in geopolymer systems can lead to the formation of hydrated calcium silicate (C-S-H) gel and Al-substituted C-S-H gel, and these are reported to reduce porosity and increase geopolymer strength [14-25]. This coexistence of geopolymer gel and hydration products has also been observed in alkali-activated fly ash and portland cement mixtures [26]. Recent research on the effect of alkali and Al on C-S-H gel confirmed that the structure is modified by alkali metals and Al [27,28]. The formation of C-S-H geopolymer or gel phase is determined by the chemical and mineralogical composition, the physical properties of aluminosilicate and the Ca source, the alkalinity of the activator and the percentage of Ca in the system [17-19,22,24]. This work builds on previous research [13] in which new geopolymers made using APC glass were described. The influence of processing parameters on the geopolymerization of APC glass was investigated. In this work, the properties of optimized APC glass geopolymers and the formation of a complex binding phase were investigated in detail. This provides new knowledge and information about the potential applications of this material and the relationship between the final properties of the geopolymer composite and the microstructure. 2. Materials and methods 2.1. Materials Glass produced by continuous plasma treatment of APC residues was supplied by Tetronics Ltd. (Swindon, UK) in the form of a coarse-grained material with Pb2+ > Ni2+ > Cd2+ and a saturation charge capacity of 0.86 mmol Cu/g. IT'S ME. El-Nahhal et al prepared a series of polysiloxane-immobilized ligand systems with amino acid functional groups through the sol-gel process and found their application in the separation of heavy metal ions from aqueous solution [14-18]. In recent years, there has been great interest in the preparation and application of functional polysilsesquioxane particles [19-21]. We recently reported the strong adsorption of Ag(I) ions on poly(3-mercaptopropylsylsesquioxane) (PMPSQ) microspheres [22]. Beari et al. [23] investigated the hydrolytic condensation of 3-aminopropyltriethoxysilane (APTES) in aqueous solutions and found that the hydrolysis and condensation products of APTES did not precipitate out of solution even after several weeks due to its excellent water solubility. In this paper, we present a one-step synthetic procedure for the synthesis of amino-functionalized polysilsesquioxane with a high content of amino groups for the development of an effective adsorbent.

X. Lu i sur. / Journal of Hazardous Materials 196 (2011) 234–241

List of SUITABLE parts b C0 Ce h k1 k2

3-aminopropyltriethoxysilane Langmuir constant (L/mmol) initial metal ion concentration (mmol/L) equilibrium metal ion concentration (mmol/L) initial adsorption rate (mmol g−1 min−1) pseudo-first order rate constant (min− 1 ) pseudo-second-order rate constant (g mmol−1 min−1) m adsorbent mass (g) MTMS methyltrimethoxysilane PAMSQ poly(aminopropyl/methyl)silsesquioxane PMSQ poly(methylsilsesquioxane) adsorption capacity (mmol/g) qm theoretical saturation adsorption capacity ( mmol/g) qt adsorption capacity at t (mmol/g) regression coefficient R2 t time (min) solution volume (L) V

heavy metals. Poly(aminopropyl/methyl)silsesquioxane (PAMSQ) particles were obtained by hydrolytic co-condensation of 3-aminopropyltriethoxysilane (APTES) with methyltrimethoxysilane (MTMS) in an aqueous medium. PAMSQ particles have the ability to effectively remove Cu(II) and Pb(II) ions from aqueous solution. The effect of adsorption time, initial concentration of metal ions and solution pH was investigated using the static adsorption method to optimize the absorbance of Cu(II) and Pb(II) PAMSQ particles. 2. Experimental 2.1. Materials 3-Aminopropyltriethoxysilane (APTES, ≥98.0%) was purchased from Diamond Advanced Material of Chemical Inc. Methyltrimethoxysilane (MTMS, ≥98.0%) was purchased from Jiangsu Danyang Organosilicon Material Industrial Corporation. Ammonium hydroxide solution (NH4OH, 25%), copper sulfate pentahydrate, and analytical grade lead nitrate were obtained commercially and used as received.

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constant temperature of 20 ◦ C at 100 revolutions per minute. After the desired adsorption period, the particles were filtered from the solution using a millipore filter membrane (0.22 µm). The final concentrations of metal ions in the solution were analyzed by atomic emission spectrometry with inductively coupled plasma (ICP-AES, IRIS 1000, Thermo Elemental). The equilibrium adsorption capacity was calculated from Eq. (1). qe =

(CO - Ce)Vm

(1)

where qe (mmol/g) is the adsorption capacity, and C0 (mmol/L) and Ce (mmol/L) are the initial and equilibrium concentration of the metal. V (L) is the volume of the solution, and m (g) is the weight of the adsorbent. 2.4. Elemental analysis measurements were performed with an elemental analyzer Vario EL III (Elementar Analysen systeme GmbH, Germany). FT-IR spectra were recorded on an iS10 FT-IR spectrophotometer (Nicolet, USA). The samples were mixed with potassium bromide and pressed into a disk to measure the absorption spectrum. High-resolution 29 Si solid-state NMR spectra were measured at room temperature on a Bruker Avance 400 MHz spectrometer (silicon frequency 99.36 MHz) equipped with a Bruker solid-state accessory. Spectra were obtained using a broadband detector head with a 4 mm magic angle double beam rotation. Chemical shifts of silicon atoms in silsesquioxane compounds are expressed using traditional Tn terminology, where the exponent corresponds to the number of oxygen bridges to other silicon atoms. Thus, the uncondensed monomer was named T0, and the fully condensed polymer without residual silanols was designated as T3 silicon atoms [24]. SEM (JSM-7401F FE-SEM, JEOL Ltd., Japan) was used to study the morphology of PAMSQ particles. The particles were coated with gold for SEM observations. Specific surface data were performed on a Micromeritics ASAP 2020 Surface and Porosity Analyzer using the BET method. TGA was performed on a thermal analyzer SDT Q600 (TA Instruments, USA) at a heating rate of 10 ◦ C min−1 in air. 3. Results and discussion 3.1. Composition and properties of PAMSQ particles

2.2. Synthesis of PAMSQ particles by hydrolytic co-condensation process PAMSQ particles were prepared using a sol-gel catalyzed base process in an aqueous medium. A mixture of APTES and MTMS in different molar ratios was added to 100 ml of water, which is 10% by weight. Ammonium hydroxide solution (0.16 mL) was added to the above solution. The reaction was continued overnight at room temperature. The resulting precipitate was then filtered through a millipore filter membrane (0.22 μm) and thoroughly washed with distilled water and ethanol several times to remove residual NH4OH as well as unreacted monomers or oligomers. Finally, the products were dried under vacuum and PAMSQ particles were obtained. 2.3. Adsorption of Cu(II) and Pb(II) on PAMSQ particles Batch adsorption experiments were carried out using PAMSQ particles as adsorbent for the adsorption of Cu(II) or Pb(II) ions from aqueous solutions of one metal ion. The pH of the sample was adjusted to the desired value with HNO3 or ammonia solution. Batch adsorption experiments were carried out in a stirred bath maintained at

The synthesis of PAMSQ particles with a controlled amount of aminopropyl functional groups was carried out using APTES and MTMS as precursors by a hydrolytic co-condensation process (Scheme 1). In general, the hydrolytic condensation of organotrimethoxysilane in water or ethanol-water was quite fast under basic conditions [23-25]. Initial hydrolysis of APTES and MTMS resulted in silanol oligomers. Silanol (Si-OH) was highly reactive and then condensed to form polysilsesquioxane in the presence of a basic catalyst. Amino groups in APTES can increase the pH of the solution and accelerate the process of hydrolytic cocondensation. Simultaneous hydrolysis of APTES and MTMS led to cocondensation, but the state of the product was different with different molar ratios of APTES/MTMS. If the molar ratio of APTES in the precursors was less than 40%, a white precipitate appeared. However, the cocondensation products did not precipitate out of solution when the molar ratio of APTES in the precursors was above 50% due to the excellent water solubility of the cocondensed PAMSQ. SEM images show the morphology of the polysilsesquioxane particles (see Figure 1). PMSQ particles prepared from MTMS alone are spherical with an average size of 2.0 µm. Particle aggregation was quite evident for PAMSQ copolymerized particles

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X. Lu i sur. / Journal of Hazardous Materials 196 (2011) 234–241

Figure 1. Schematic representation of the synthetic strategy for PAMSQ particles. Note that "OR" can contain silanol (OH) or other silane units

and the particle size decreased with increasing amount of APTES in APTES/MTMS mixtures, as shown in the figure. 1b-d. The results may be due to the excellent solubility of APTES in water, which makes the copolymerized PAMSQ particles more hydrophilic. In addition, the amino groups in APTES increased the pH of the reaction solution and catalyzed the hydrolytic co-condensation reactions of APTES and MTMS. An increase in the rate of hydrolysis with increasing pH resulted in a higher nucleation rate, which also resulted in a larger number of particles but a smaller final particle size [24]. Table 1 shows the properties of polysilsesquioxane particles. The content of amino groups of PAMSQ particles determined by elemental analysis was lower than the theoretical values. The results may be due to the excellent solubility of APTES and its hydrolysis and condensation products in water [23]. These results are consistent with those reported by Liu et al. [20] copolymerized aminopropyl/phenylsilsesquioxane microparticles synthesized by hydrolytic co-condensation of APTES and phenyltriethoxysilane (PTES). The specific surface areas of polysilsesquioxane particles were estimated using the BET method as shown in Table 1.

Relatively small values of the BET surface area were due to the complete condensation of polysilsesquioxane. The effect of the content of amino groups on the adsorption of Cu(II) and Pb(II) on PAMSQ particles was investigated. As shown in Table 1, the aminopropylated PAMSQ samples showed a high affinity for Cu(II) and Pb(II), and the adsorption capacity of metal ions on PAMSQ particles increases with the increase in the content of amino groups. However, unmodified PMSQ without amino groups adsorbed only a small amount of Cu(II) and Pb(II) ions. The results in Table 1 show that the adsorption mechanism primarily involves the complexation of metal ions with amino groups. The high content of amino groups makes PAMSQ3 suitable as a representative copolymer product for further research in adsorption experiments. Solid-state 29 Si NMR spectra are powerful methods for characterizing the chemical structure of polysilsesquioxane structures. According to Arkhireeva et al. [26], in the case of MTMS-derived polysilsesquioxane, the resonances at -65.9, -57.1, -48.5 ppm can be attributed to the T3, T2, and T1 species, respectively. And Caravajal et al. [27] published 29Si NMR spectra

Fig. 1. SEM images of polysilsesquioxane particles: (a) PMSQ, (b) PAMSQ1, (c) PAMSQ2 and (d) PAMSQ3.

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237

Table 1. Properties of polysilsesquioxane particles. Samples

APTES i komonomer (mol%)

Amino acid content in PAMSQ (mol%)

BET surface area (m2/g)

b Cu(II) adsorption capacity (mmol/g)

b Pb(II) adsorption capacity (mmol/g)

PMSQ PAMSQ1 PAMSQ2 PAMSQ3

0 20 30 40

0 15 22 27

4,3 4,8 6,7 5,0

0,13 0,80 1,62 2,25

0,17 0,58 0,90 1,14

a b

The amino acid content of PAMSQ was determined by elemental analysis of the nitrogen content. Initial concentration of metal ions: 10 mM, adsorption time: 5 h, adsorbent dose: 2 g/L.

1482 cm-1 (ıN-H), 3376 cm-1 (N-H), 2934 and 2970 cm-1 (C-H). The broad band around 3400 cm-1 can be attributed to both adsorbed water and the Si-OH group [29]. The thermal stability of PMSQ and PAMSQ3 in air was investigated by TGA (Figure 4). The weight loss in the range of 100–250 ◦ C for polysilsesquioxane is probably due to the residual reaction of the alkoxysilyl groups [26]. It was found that the thermal reduction of polysilsesquioxane in the range of 250-700 ◦ C is mainly due to the decomposition of organic parts [24]. Thermal decomposition temperatures of PAMSQ3 particles in air at 5% weight loss and 10% weight loss are 257 ◦ C and 370 ◦ C, respectively. The TGA result shows that both PMSQ particles and PAMSQ3 particles have good thermal stability.

Fig. 2. Steady state

Si NMR spectrum of PAMSQ3 particles.

3.2. Adsorption kinetics of Cu(II) and Pb(II) on PAMSQ particles

showed large peaks in the regions of -66, -58, and -49 ppm due to silicon in the bound CH2CH2CH2NH2 residue of the APTES-modified silica. Fig. 2 shows the solid 29Si NMR spectrum of PAMSQ3 particles. There is a large peak at -62.1 ppm and a weak peak at around -53 ppm which is attributed to fully condensed T3 or linear T2 type. The formation of T1 and T0 species is negligible, suggesting that cocondensation is complete [28]. Fig. Fig. 3 shows the IR spectra of PMSQ and PAMSQ3 particles. PMSQ and PAMSQ3 show well-defined methyl group and Si-O-Si absorption bands at: 2970, ca. 2921-2934 cm-1 (C-H), 1410 cm-1 (ıC-H in Si-R), 1272 cm-1 (ıC-H in Si-R), approx. 1119-1127, approx. 1032-1036 cm-1 (Si-O-Si) and 778 cm-1 (Si-C) [29,30]. The Si-O-Si stretching peaks at 1119-1127 cm-1 indicate the presence of a cage structure, while the absorption at 1032-1036 cm-1 indicates that the ordered structure is probably ladder-like or layered [20,30] . The spectrum of PAMSQ3 showed bands due to aminopropyl groups at

Adsorption kinetics is one of the important properties that determine adsorption efficiency. The effect of contact time on the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles is shown in the figure. 5. The kinetic curve shows that the adsorption was fast in the first 10 minutes, when the adsorption capacity reached 1.49 mmol/g and 0.58 mmol/g for Cu(II) and Pb(II), respectively, and then gradually slowed down. The initial fast step of metal ion adsorption can be attributed to physical and reactive adsorption between metal ions and amino groups on the surface of PAMSQ particles. However, the next slow step is attributed to intraparticle adsorption, which represents the diffusion of Cu(II) and Pb(II) ions inside the particles over a long time. Experimental results suggest that the amount of adsorbed metal ions increased with increasing adsorption time and reached equilibrium after 300 min for Cu(II) and Pb(II). Therefore, in this study, we used a contact time of 300 minutes for further experiments. Pseudo-first-order and pseudo-second-order models [31] were used to test the experimental data and thus clarify

Fig. 3. IR spectra of PMSQ and PAMSQ3 particles.

Figure 4. TGA curves for PMSQ and PAMSQ3 in air.

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X. Lu i sur. / Journal of Hazardous Materials 196 (2011) 234–241

Fig. 5. Effect of adsorption time on Cu(II) and Pb(II) adsorption on PAMSQ3 particles (initial concentration of metal ions: 10 mM, adsorbent dose: 2 g/L).

kinetic adsorption process. Lagergren pseudokinetic model of the first order, represented as: k1 log(qe − qt ) = log qe − t 2.303

(2)

where qe and qt are the amounts of adsorbed metal ions (mmol/g) at equilibrium and time t, respectively, and k1 (min−1 ) is the pseudo-first-order rate constant. qe and the rate constant k1 were calculated by plotting log (qe − qt) against t. As can be seen in fig. 6, the pseudo-first-order model does not fit the data well. The experimental and calculated qe values, pseudo-first-order rate constants and regression coefficient (R2) values are shown in Table 2. The calculated qe values in the pseudo-first-order model did not agree with the experimental qe values, suggesting that the adsorption of Cu(II) and Pb(II) does not follow pseudo-first kinetics. In order to find a more reliable description of the adsorption kinetics of Cu(II) and Pb(II) ions on particles, a second-order pseudokinetic model was applied to the experimental data. The pseudo-second-order equation can be written as: t 1 1 = + t qt qe k2 q2e

(3)

where qe and qt are defined as in the first-order pseudokinetic model. k2 is the pseudo-second order rate constant. The slope and intercept of t/qt versus the t line of the graph in Figure 7 gave the values

Fig. 6. Pseudokinetic diagrams of the first order for the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles.

Fig. 7. Pseudo-second-order kinetic diagrams for Cu(II) and Pb(II) adsorption on PAMSQ3 particles.

of qe and k2. Additionally, the initial adsorption rate (h) can be determined from the values of k2 and qe using h = k2 q2e. The regression coefficients (R2) and several parameters obtained from the pseudo-second-order kinetic model are also shown in Table 2. As can be seen from Table 2, the calculated values of qe are in good agreement with the experimental values of qe. Additionally, the obtained R2 values for Cu(II) and Pb(II) adsorption are both above 0.99. Therefore, the adsorption kinetics could be better approximated by the pseudo-second-order kinetic model for Cu(II) and Pb(II) on PAMSQ particles. A pseudo-second-order model was developed based on the hypothesis that the rate-determining step can be chemisorption driven by covalent forces through electron exchange or valence forces through adsorbate-adsorbate electron sharing [31], indicating that the adsorption of Cu( II ) and Pb (II ) on PAMSQ3 particles is mostly chemically reactive adsorption. 3.3. The effect of the initial concentration of metal ions and the adsorption isotherm The effect of the initial concentration of metal ions on the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles is shown in the figure. 8. At a lower initial concentration of metal ions, abundant aminopropyl groups on the surface of PAMSQ particles can react with metal ions, resulting in significantly increased adsorption of Cu(II) and Pb(II). Then, the adsorption process gradually slows down with an increase in the initial concentration of metal ions.

Fig. 8. Effect of initial concentration of metal ions on the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles (initial concentration of metal ions: 1–20 mM, adsorption time: 5 h, adsorbent dosage: 2 g/L).

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239

Table 2 Kinetic model equations for Cu(II) and Pb(II) adsorption on PAMSQ3 particles. Metal ions

Cu(II) Pb(II)

q (nastavak) (mmol/g)

2,25 1,14

Pseudo-first class

Pseudo-second order

k1 (min-1)

qe (kal.) (mmol/g)

k2 (g mmol−1 min−1)

h (mmol g−1 min−1 )

qe (kal.) (mmol/g)

0,01073 0,01262

1,37 0,85

0,8695 0,8830

0,021 0,028

0,11 0,04

2,30 1,22

0,9915 0,9908

a facile method under modest conditions would have promising applications as a cost-effective adsorbent. 3.4. Effect of solution pH on adsorption

Fig. 9. Langmuir diagrams of Cu(II) and Pb(II) adsorption on PAMSQ3 particles (initial concentration of metal ions: 1–20 mM, adsorption time: 5 h, adsorbent dose: 2 g/L).

Fig. Figure 10 shows the effect of solution pH on the adsorption of Cu(II) and Pb(II) by PAMSQ3 particles. A pH in the range 2.0-5.0 was chosen to avoid precipitation of Cu(OH)2 and Pb(OH)2. The adsorption capacity increased with increasing pH of the solution in the range of pH 2.0–5.0, and no adsorption was observed at pH 2.0. This can be attributed to competitive adsorption between metal ions and H + ions on PAMSQ3 particles. At low pH, the adsorption of metal ions decreases because high concentrations of competing H + ions occupy adsorption sites, while protonated amino groups are deprotonated with increasing pH, increasing the absorbance of metal ions [15,44]. Therefore, the pH of the solution around 5.0 could be optimal for the application of PAMSQ3 particles as an effective Cu(II) and Pb(II) adsorbent. 3.5. Mechanism of adsorption of metal ions on PAMSQ particles

Figure 9 shows the adsorption isotherms for Cu(II) and Pb(II) PAMSQ3 particles at 20 ◦ C. The adsorption data are plotted according to the Langmuir equation: Ce 1 Ce = + qe qm qm b

(4)

where qm and b are characteristic Langmuir parameters. qm is the theoretical saturation adsorption capacity of the monolayer (mmol/g), and b is a constant associated with the intensity of adsorption. Plotting Ce /qe against Ce gives straight lines as shown in fig. 9. Table 3 shows the coefficients of the Langmuir model along with the regression coefficients (R2). As can be seen from Table 3, the R2 values of the Langmuir isotherm models were above 0.99, indicating that the Langmuir model fits the experimental results well. The calculated values of qm are in good agreement with the experimentally determined ones. PAMSQ3 particles have a strong ability to adsorb Cu(II) and Pb(II) ions from aqueous solutions, which indicates a great potential as a high efficiency adsorbent. Variations in the intake of metal ions on different adsorbents are related to adsorbent properties such as structure, functional groups and specific surface area. Table 4 shows a comparison of the maximum adsorption capacity of PAMSQ for Cu(II) and Pb(II) on different adsorbents mentioned in the literature. The results show that the adsorption capacity of PAMSQ3 particles for Cu(II) and Pb(II) was high compared to many other adsorbents. Therefore, it can be believed that the PAMSQ3 particles were synthesized from a common silane coupling agent via

Sorption is generally defined as the transfer of ions from the solution phase to the solid phase through various mechanisms such as physical and chemical adsorption, surface deposition or diffusion, or fixation in the solid state [45]. According to Pearson's theory of hard and soft acids and bases [11], PAMSQ activated by the aminopropyl group has the ability to bind to heavy metal ions such as Cu(II) and Pb(II). FTIR spectra of PAMSQ3 particles before and after adsorption of Cu(II) ions are shown in the figure. 11. The appearance of a sharp peak at 619 cm−1 after the adsorption of Cu(II) on PAMSQ3 is attributed to the stretching vibration of the N – Cu bond formed during the complexation process [37]. FTIR results confirm that nitrogen in PAMSQ3 particles actively participates during the adsorption process through complexation with Cu(II) ion. Therefore, the mechanism of adsorption of metal ions on PAMSQ3 particles primarily involves the complexation of metal ions with amino groups.

Table 3 Langmuir isotherm coefficients. Metal ions

qm (mmol/g)

qm (kal.) (mmol/g)

b (L/mmol)

Cu(II) Pb(II)

2,29 1,31

2,26 1,47

11,82 0,48

0,9986 0,9925

Fig. 10. Effect of pH on the adsorption of Cu(II) and Pb(II) on PAMSQ3 particles (initial metal ion concentration: 2.5 mM, adsorption time: 5 h, adsorbent dosage: 2 g/L).

240

X. Lu i sur. / Journal of Hazardous Materials 196 (2011) 234–241

Table 4. Comparison of the maximum adsorption capacity of PAMSQ for Cu(II) and Pb(II) on different adsorbents mentioned in the literature. Adsorbents

Adsorption capacity (mg/g)

SBA-15 mezoporozni silicij s dendrimernim aminima na bazi melamina xantat Aminirana polyakrilonitrilanna nanovlakna Porozni chitosan monoliti PS-EDTA smola 2-(((2-aminoetilamino) kooksidirani karcino-amino) fuginirani metil-aktivirani) ludge Ulva lactuca alge Kalijev hydroxyd -tretirani prah šišarki PAMSQ

Cu(II)

Pb(II)

Ref.

126 19,8 1,21 99,4

130 16,8 1,54

[32] [33] [34] [35] [36] [16] [37] [38] [39] [40] [41] [42] [9] [43] This project

80,19 130,91 43,47 116,52 141,8 42,1 12,1 17,3 112 19,22 146

32,1 16,2 42,4 230 26,27 272

Fundamental Research Funds for Central Universities (Project No. WA1013012) and East China University of Science and Technology (ECUST) gennem Promotion of Undergraduate Innovative Experimental Program (No. X0807).

bibliographical references

Fig. 11. FTIR spectra of PAMSQ3 particles (a) before adsorption and (b) after Cu(II) adsorption. Inset: copper ion binding scheme.

4. Conclusions Amino-functional polysilsesquioxane particles were synthesized by hydrolytic co-condensation using APTES and MTMS as precursors in the presence of a base catalyst in an aqueous medium. The process is a one-step cocondensation synthesis route, where the functionalities of the particles can be easily controlled by changing the organosilane feed ratio. The results of solid state NMR spectroscopy, FT-IR analysis and elemental analysis confirmed co-condensation between organosilanes. PAMSQ particles have been shown to be an effective adsorbent for the removal of Cu(II) and Pb(II). The adsorption behavior of Cu(II) and Pb(II) on PAMSQ particles is influenced by the adsorption time, the initial concentration of metal ions and the pH of the solution. Kinetic studies showed that the adsorption process fits well with pseudo-second-order kinetics with a high initial adsorption rate. The experimental data fit the Langmuir isotherm model well. PAMSQ particles show the highest Cu(II) and Pb(II) adsorption capacities of 2.29 mmol/g and 1.31 mmol/g, respectively, at the initial metal ion concentration of 20 mM. Therefore, there are good prospects for PAMSQ particles in practical applications for the removal of Cu(II) and Pb(II) ions from their aqueous solutions. Acknowledgments This work was financially supported by the National Natural Science Foundation of China (Project No. 21006025),

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The rest of the snow was covered by snow on snow , Domenico Pirozzi b , Serena Esposito c , Filomena Sannino a,∗ a Department of Soil, Plant, Environmental and Animal Production Sciences, Faculty of Science, University of Science "Federico II", Via Università 100, 80055 Portici, Naples, Italy b Department of Chemical Engineering, Faculty of Engineering, University of Naples "Federico II", P.le Tecchio, 80, 80125 Naples, Italy c Materials Laboratory at the Department of Mechanics, Structures, Environment and Territory, Technical faculty of the University of Cassino, Via G. di Biasio 43, I-03043, Cassino (Fr), Italy

article

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Article history: Received April 19, 2011 Received in revised form September 6, 2011 Accepted September 6, 2011 Available online September 12, 2011 Keywords: Decontamination Simazine Adsorption Mesoporous metal oxides Regeneration

a b s t r a c t Two mesoporous metal oxides, Al2O3 and Fe2O3, were evaluated for their ability to remove simazine, a highly persistent s-triazine herbicide, using a balanced batch method. The effect of several experimental parameters such as pH, contact time, initial concentration and sorbent dose on herbicide sorption was investigated. The maximum absorption of simazine on Al2O3 and Fe2O3 was observed at pH 6.5 and 3.5, respectively. The different adsorption capacities of the two oxides are explained by considering a set of factors that influence the adsorption process, such as surface area and porosity. Adsorption kinetics on both oxides were described using a pseudo-second-order model. Sorption of simazine on Fe2O3 was faster compared to that observed on Al2O3. Alumina has been shown to be regenerated by incineration and therefore may be considered for industrial treatment systems designed to mitigate pesticide contamination in the aquatic environment. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Water pollution by pesticides has been recognized in agricultural areas of the world for many years, and abundant evidence suggests that many water resources are contaminated with organic pesticides. Common agricultural practices, accidental spills or uncontrolled release of contaminated water due to washing of pesticide containers or industrial wastewater into the environment result in pollution of air, soil, surface and underground water and living organisms. Due to the protection of the environment and human health, it is therefore important to develop new rehabilitation technologies. Adsorption is currently believed to be a simple and effective technique for water and wastewater treatment, and its success is largely dependent on the development of effective sorbents. Activated carbon [1], clay minerals [2], biomaterials [3], zeolites [4] and some industrial solid wastes [5] have been widely used with varying efficiency. In the wastewater treatment process that uses sorption,

∗ Corresponding author at: Department of Soil, Plant, Environmental and Animal Science, University of Naples Federico II, Via Università 100, 80055 Portici (NA), Italy. Phone: +39 081 2539187/2539183; fax: +39 081 2539186. E-mail address:[email protected](F. Sannino). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.022

sorbent regeneration is necessary. However, the high costs associated with the regeneration of sorbents or the necessity of extraction obtained with acidic or alkaline solutions represent a serious problem. New sorbents are needed to remove organic pollutants in water purification processes. An ideal sorbent should have a large surface area (ie, a high density of adsorption sites), evenly accessible pores, and physical and/or chemical stability [6]. The adsorption capacity of a sorbent is believed to be largely determined by its surface area, which usually increases with decreasing particle size, although the pore size distribution is also critical for an optimal adsorption process. Therefore, thanks to the introduction of nanostructured oxide materials, the efficiency of pollutant removal can be dramatically increased. Mesoporous materials, a class of nanoporous materials, have attracted much attention in the scientific and industrial community since the introduction of well-ordered mesoporous silicas that have large surface areas and uniform and tunable pore sizes (2–50 nm). 7,8]. The great interest of these materials as adsorbents for environmental remediation is not only due to their large surface area, but also due to the fast kinetics of pollution adsorption. Recent works [9-11] have shown that mesoporous materials can have high adsorption capacity, good selectivity and improved recovery ability for the removal of toxic compounds from aqueous solutions.

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The encouraging results obtained from these studies led us to investigate the sorption of simazine (2-chloro-4,6bis(ethylamino)-s-triazine), a key herbicide from the strazine family, onto mesoporous metal oxides. s-triazines are selectively resistant herbicides that have been widely studied due to their increasingly widespread use in forestry and pre- and post-emergence in agricultural soils [12]. Although these herbicides are now banned in some countries, the resistance of s-triazines to chemical and biological degradation has led to their accumulation in the environment [12]. In Italy, Annex 5, included in Legislative Decree 152/2006 on the environment, states the safe limit of atrazine (s-triazine herbicide) in soil from 0.01 to 1.0 mg kg−1, while in water it is 0.3 ␮g L−1. Simazine is a synthetic s-triazine herbicide used to control broadleaf weeds and annual grasses in agricultural and non-agricultural fields [13,14]. It is the second most frequently detected pesticide in surface and groundwater in the USA, Australia and Europe [15]. Due to the carcinogenic potential of s-triazine, the presence of simazine in water is of increasing concern [16]. A significant amount of research has been conducted on the removal of s-triazine by sorption to soil and to various organic and inorganic sorbents [17,18]. However, to our knowledge, no work has been published on the sorption capacity of mesoporous oxides towards triazine. Therefore, the aim of this work was to evaluate two commercial metal oxides with a mesoporous structure (Al2O3 and Fe2O3) regarding their ability to remove simazine from aqueous solutions. In view of future applications, the regeneration of these materials is also discussed.

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2.3. Sorption test The basic herbicide solution was prepared by dissolving 2 mg of simazine in 500 ml of 0.03 M KCl (final concentration 20 µmol L-1). This solution was then stored in a refrigerator. All sorption experiments were performed by adding 10 mg of sorbent to 20 ml of simazine solution in glass vials with Teflon stoppers at a temperature of 20 ◦ C. After incubation for 24 hours on a rotary shaker (35 revolutions per minute), the samples were centrifuged at 7000 rpm 20 min. The supernatant was analyzed to estimate herbicide concentration using a high-pressure liquid chromatography (HPLC) technique as described below. The amount of simazine adsorbed on the oxides was calculated as the difference between the initial amount of added herbicide and that present at equilibrium. Blanks of simazine in 0.03 M KCl were analyzed to verify pesticide stability and adsorption onto vials. Several experiments were conducted to study the effect of various factors affecting the sorption of simazine onto Al2O3 and Fe2O3, as summarized below:

2-Chloro-4,6-bis(ethylamino)-1,3,5-triazine (simazine) (Figure S1 in the Supporting Information ) was purchased from Sigma-Aldrich Chemical Company (Poole, Dorset, UK, purity 99.0 % ). All solvents were of HPLC grade (Carlo Erba, Milan, Italy) and were used without further purification. The water used to prepare all solutions was obtained from a Millipore Waters Milli-Q water purification system. All other chemicals were obtained from Sigma-Aldrich unless otherwise noted. ␥-aluminum (Al2O3) and iron(III) (Fe2O3) oxides were purchased from IoliTec Nanomaterials (Denzlingen, Germany; 99.9 and 99.5% purity for Al2O3 and Fe2O3, respectively).

(a) Effect of pH: Experiments were carried out by adding pesticide solutions at a constant concentration (10 ␮mol L−1) and different pH values from 3.0 to 7.0. The pH was controlled by the addition of 0.10 or 0.01 mmol L-1 HCl or KOH solution. The samples were shaken for 24 hours and then, after centrifugation, analyzed as described below. (b) Effect of amount of sorbent: Experiments were carried out by adding pesticide solutions in two concentrations (5 and 10 ␮mol L−1), in different solid/liquid ratios. Ratios of 0.1, 0.5, 1.0 and 2.0 were obtained by adding 2.0, 10, 20 and 40 mg of Al2O3 and Fe2O3, respectively, to a final volume of 20 mL at 20 ◦ C. The samples were incubated at pH 65 (test with Al2O3) and 3.5 (test with Fe2O3), during 24 hours. (c) Effect of incubation time: Kinetic studies were performed using 10 ␮mol L-1 simazine solution at pH 6.5 (test with Al2O3) and pH 3.5 (test with Fe2O3). The solutions were mixed for 2.0, 5.0, 20, 40, 60, 90, 120, 320, 960 and 1800 min. (d) Sorption isotherm: Different volumes of herbicide stock solution (20 µmol L-1) were added to each oxide to give an initial concentration of simazine ranging from 0.50 to 10.69 ␮mol L-1. The pH of each solution was kept constant at pH 6.5 (test with Al 2 O 3 ) and 3.5 (test with Fe 2 O 3 ) by the addition of 0.10 or 0.01 mmol L-1 of HCl or KOH solution. The samples were incubated for 20 minutes (Al2O3 test) and 180 minutes (Fe2O3 test). Then, after centrifugation, the supernatants were analyzed as described below.

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2.2. Chemical and physical analysis of Al2O3 and Fe2O3

2.4. Detailed provision

The determination of the point of zero charge (pzc) for Al2O3 and Fe2O3 was carried out according to the methods described by Addorisio et al. [11]. The specific surface area (SSA) of Al2O3 and Fe2O3 was calculated using the Brunauer-Emmett-Teller (BET) method [19]. N2 adsorption-desorption isotherms at 77 K were obtained with a Micromeritics Gemini II 2370 instrument. Before each measurement, the sample was degassed at 250 ◦ C for 2 h under a stream of N2. The pore volume was determined from the amount of adsorbed N2 at P/P◦ = 0.98 (desorption curve), assuming the presence of liquid N2 (density = 0.807 g cm−3) in the pores under these conditions. The average values for the pore diameter dp were calculated from the relationship: dp = 4V/ABET, where V is the total pore volume. The Barrett-Joyner-Halenda (BJH) approach [19] was used to calculate the sample pore size distribution using desorption data.

Simazine was analyzed with an Agilent 1200 Series HPLC instrument (Wilmington, USA), equipped with a DAD array and Agilent ChemStation software. The analysis procedure is described in detail in the Supporting Information.

2. Materials and methods 2.1. Materials

2.5. Diffuse Reflectance Fourier Transform Spectroscopy (DRIFTS) Analysis The sample preparation procedure for DRIFTS determination is described in detail in the Supporting Information. 2.6. Scanning Electron Microscopy (SEM) Analysis SEM analysis of the Al2O3 samples at pH 4.0 and 6.5 was performed by an FEI Quanta 200 ESEM.

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-1

Simazin sorbate (μmol kg)

Al2O3

Fe2O3

6000 5000 4000 3000 2000 1000 0

3.0

4.0

5,0

6,0

7,0

8,0

Fig. 1. Effect of pH on the adsorption of simazine by Al2O3 and Fe2O3 at a solid/liquid ratio of 0.5.

2.7. Data analysis All experiments were performed in triplicate and the relative standard deviation was in all cases less than 3%. 3. Results and discussion 3.1. Effect of pH In order to estimate the optimum pH to be used in subsequent experiments, sorption experiments were carried out to examine the effect of pH at a solid/liquid ratio of 0.5, which was provisionally determined to be the optimum value for sorption. The results shown in Fig. Figure 1 shows that the highest sorbed amount of simazine was observed at pH 6.5 when Al 2 O 3 was used and at pH 3.5 in tests with Fe 2 O 3 . In aqueous solution, triazines such as simazine exist in a neutral or protonated form, depending on the pKa of the compound (the pKa of simazine is 1.70) and the pH of the system. The ring nitrogen, located at position 3 between the electron-rich alkyl side chains, is the most basic and therefore the most likely protonation site. At low pH values (e.g. 3.0-3.5) the surfaces of oxides and soluble species are strongly protonated [20], so that the more basic triazine nitrogen (N-3) can easily form a bond resonance with Fe2O3, which results from overlapping lone a pair of nitrogen electrons and partially filled metal d orbitals in iron, the latter being a transition metal (d-block element). Adsorption of simazine on Al2O3 as a function of pH is quite different. A possible explanation of our results is that the key role in herbicide sorption is played by the oxidation state, which is greatly influenced by the pH value of the medium. Specifically, as shown in FIG. 2, large aggregates and small particles were observed at pH 4.0 and pH 6.5, respectively. Accordingly, the textural properties of Al2O3 are expected to be modified by pH. In order to confirm the latter hypothesis, physical analysis was performed on the Al2O3 sample at pH 4.0 and pH 6.5 by analyzing the relative N2 adsorption-desorption isotherms. As shown in Table 1, the surface area of the Al2O3 sample at pH 4.0 (157 m2 g-1) can be compared with that at pH 6.5 (150 m2 g-1). Table 1. Physical properties of Al2O3 at pH 4.0 and pH 6.5. Sample

ABET (m2 g−1 )

Pore volume (cm3 g−1 )

Mean dp (nm)

Al2O3 (pH 4,0) Al2O3 (pH 6,5)

157 150

0,352 0,643

8,9 17,3

Fig. 2. SEM image of Al2O3 at 500x magnification at pH 4.0 (a) and at pH 6.5 (b).

However, more interesting clues can be obtained from the comparison between the pore size distributions (Table 1), obtained by processing the desorption data using the BJH method. It is clear that the contact with solutions at different pH values greatly affects the internal organization of the Al2O3 particles, creating the porosity of smaller voids in the case of the Al2O3 sample at pH 4.0. In addition, the pore volume of the samples at pH 6.5 is much larger than that observed at pH 4.0 (Table 1). These observations were confirmed by SEM analysis (Figure 2). Finally, at pH 6.5, the herbicide could give an acid-base reaction with Al2O3, which at this pH is present as Al[(H2O)6]3+. Alternatively, since simazine is more nucleophilic than water molecules, it may replace some water molecules in the hexacoordinate complex. Chappell et al. [21] demonstrated the interaction of atrazine with smectite surfaces through hydrogen bonding and simultaneously showed that the alkyl tails of herbicides can interact with hydrophobic nanosites on smectite basal surfaces. Other studies have shown that non-covalent binding forces, cation-␲, can occur between s-triazine and a metal cation [22]. In the literature, data on the binding mechanism of triazine herbicides to oxides are very scarce. Accordingly, the explanation of the above observed behavior deserves special attention.

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245

-1

Simazin sorbate (μmol kg)

AlO2

1,2 104 1 10

Fe O

8000 6000 4000 2000 0

Equilibrium concentration of simazine (μmol L) -1

Fig. 3. The effect of the solid/liquid ratio on the adsorption of two different concentrations of simazine with Al2O3 and Fe2O3 at pH 3.5 and 6.5, respectively.

Alternatively, kinetic adsorption curves were analyzed using a second-order pseudokinetic model:

3.2. Effect of solid/liquid ratio Simazine sorption studies were conducted using Al2O3 (at pH 6.5) and Fe2O3 (at pH 3.5), varying the amount of sorbent and adding two different herbicide concentrations. The results shown in Fig. Figure 3 shows, for both oxides and regardless of herbicide concentration, greater adsorption at a solid/liquid ratio of 0.5. In particular, the amount of adsorbed herbicide on Al2O3 was already significant at the lowest solid/liquid ratio (0.1) and increased significantly with increasing amount of oxide. However, no sorption of simazine was observed in the presence of 20 and 40 mg of oxide. It is clear that the greater the amount of oxide, the greater the diffusion resistance in the mesoporous structure, which leads to less herbicide absorption.

3.3. Effect of incubation time Kinetic data were analyzed using the pseudokinetic equation of the first order [23]: dq = k1 · (qe − q) dt

(E1)

where qe and q are the amounts of adsorbed herbicide (␮mol kg−1) at equilibrium and time t, respectively, k1 is the sorption rate constant (min−1) and t is time (min). Integration with the boundary condition q|t=0 = 0 Eq. (E1) the following expression is obtained: log(qe − q) = log qe −

At t 2,303

0,20

(E2)

AlO2

Fe O 2

-1

t/q (min. kg µmol)

sl. 5. Isoterminal sorpcije simazina s Al2O3 i Fe2O3.

0,15

dq = k2 · (qe − q)2 dt

(E3)

where k2 is the adsorption rate constant (kg ␮mol−1 min−1). By integrating with the boundary condition q|t=0 = 0, Eq. (E3) gives the following expression: t 1 t − = q qe k2 · q2e

(E4)

The best model to describe the kinetic sorption data was a pseudo-second-order model (ie, Eq. (E3)), as shown by the linear behavior of the (t/q) vs. time plot (Fig. 4). The corresponding model parameters (qe and k2) were evaluated with respect to simazine sorption on Al2O3 (qe = 6098 ␮mol kg−1, k2 = 2.85 × 10−5 kg ␮mol−1 min−1, r2 = 0.99 ), and in Fe2O3 (qe = 1695 ␮mol kg−1, k2 = 9.03 × 10−4 kg ␮mol−1 min−1, r2 = 0.99). Sorption on Fe2O3, which reached equilibrium after 5 minutes, was faster compared to that on Al2O3, showing an equilibrium time of 120 minutes. Therefore, all equilibrium determinations were performed using an incubation period of 20 min for Fe2O3 and 180 min for Al2O3. 3.4. Sorption isotherm Sorption isotherms for simazine on Al2O3 and Fe2O3 are shown in Fig. 5. The obtained data were analyzed according to the Freundlich equation: x = Kc 1/N

(E5)

where x is the amount of absorbed pesticide (␮mol kg−1 ), cequilibrium pesticide concentration (␮mol L−1), K [(␮mol kg−1 )/(␮mol L −1 )1/ N ] and N (dimensionless) are constants that give estimates of sorption capacity or intensity, according to Giles et al. [24]. Simazine sorption isotherms on Al2O3 and Fe2O3, shown in Fig. 5, are well adapted to the linear form of the Freundlich equation (r2 > 0.99) (Table 2). According to the classification of Giles et al.

0.10 Table 2 Freundlich parameters for simazine sorption on Al2O3 and Fe2O3.

0,05

Friendly parameters

100

150

Time (min) Fig. 4. Effect of time on the sorption of simazine Al2O3 and Fe2O3.

200

Al2O3Fe203a

K (␮mol kg−1)/(␮mol L−1)1/N

N (no dimensions)

r2 a

168,11 156

0,44 0,56

0,99 0,99

Correlation coefficient.

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Table 3. Comparison of surface area, pore volume and mean pore diameter for Al2O3 and Fe2O3 samples. Sample Al2 O3 Al2 O3 500 Fe2 O3 Fe2 O3 500

ABET (m2 g−1) 195 200 106 33

Pore volume (cm3 g−1 )

Mean dp (nm)

0,723 0,770 0,239 0,0650

14,8 14,7 9,2 7,9

[24], οι πειραματικές ισόθερμες προσρόφησης ήταν τύπου S για το Al2O3 και τύπου C για το Fe2O3. Συγκεκριμένα, σε χαμηλές συγκεντρώσεις ισορροπίας η απορροφούμενη ποσότητα σιμαζίνης στο Fe2 O3 ήταν παρόμοια με εκείνη που ανιχνεύθηκε στο Al2 O3, ενώ σε συγκεντρώσεις μεγαλύτερες από 3 ␮mol L−1 υπήρχε μια αξιοσημείωτη διαφορά στη συμπεριφορά των δύο μεσοπορωδών οξειδίων. Στην πραγματικότητα, σε μια συγκέντρωση ισορροπίας 6,0 ␮mol L−1 σιμαζίνης, η ποσότητα του ζιζανιοκτόνου που προσροφήθηκε στο Fe2O3 ήταν 4000 ␮mol kg−1, ενώ αυτή που προσροφήθηκε στο Al2O3 ήταν ∼8000 ␮mol kg−1. Η ισόθερμος τύπου S της σιμαζίνης στο Al2 O3 υποδεικνύει ότι η παρουσία μορίων ζιζανιοκτόνου που έχουν ήδη απορροφηθεί στην επιφάνεια ευνοεί τη διαδικασία ρόφησης με συνεργατικό αποτέλεσμα. Αυτό το φαινόμενο μπορεί να εξηγηθεί υποθέτοντας ότι τα μόρια που έχουν ήδη απορροφηθεί τροποποιούν τη συγγένεια των θέσεων προσρόφησης προς τα μόρια που υπάρχουν στο διάλυμα. Αντίθετα, η ισόθερμος τύπου C της σιμαζίνης στο Fe2O3 χαρακτηρίστηκε από μια ευθεία τάση, ενδεικτική μιας σταθερής κατανομής του ζιζανιοκτόνου μεταξύ διαλύματος και ροφητή μέχρι να φτάσει σε κορεσμό. Οι σταθερές Freundlich (Κ και Ν) (Πίνακας 2) έδειξαν ότι το Al2O3 προσρόφησε το ζιζανιοκτόνο με υψηλότερη απορροφητική ικανότητα και χαμηλότερη συγγένεια σε σύγκριση με το Fe2O3. Η παρουσία δευτερογενών μικρών πόρων στο όριο της περιοχής μικροπόρων στο Al2 O3 μπορεί να επηρεάσει θετικά την απορρόφηση μικρών οργανικών μορίων όπως η σιμαζίνη (0,784 nm), καθώς είναι πιθανό η ενέργεια ρόφησης να αυξάνεται σε εκείνους τους πόρους των οποίων οι διαστάσεις πλησιάζουν τις διαστάσεις του ζιζανιοκτόνου (0,7–0,9 nm). Πράγματι, όπως αναφέρεται στο Σχήμα S2 των Υποστηρικτικών Πληροφοριών, η κατανομή μεγέθους πόρων του Al2O3 φαίνεται να είναι διτροπική, που χαρακτηρίζεται από δύο μέγιστα στα περίπου 3 nm και στα περίπου 15 nm. Αντίθετα, το Fe2O3 παρουσιάζει μονοτροπική κατανομή και το μεγαλύτερο μέρος του όγκου N2 προσροφάται στην περιοχή μεγέθους πόρων 6-10 nm [11]. Ένας συνδυασμός παραγόντων, όπως το εμβαδόν επιφάνειας και το πορώδες, επηρεάζει σημαντικά την υψηλότερη ικανότητα ρόφησης του Al2 O3 από το Fe2 O3 (βλ. Πίνακα 3). Τέλος, πραγματοποιήθηκαν αναλύσεις DRIFT σε κάθε οξείδιο μετάλλου μετά την ρόφηση του ζιζανιοκτόνου. Τα φάσματα DRIFT των συμπλοκών Al2O3 και Fe2O3-σιμαζίνης καταγράφηκαν και συγκρίθηκαν με εκείνα της σιμαζίνης και των μη επεξεργασμένων οξειδίων (Σχήμα S3 των Υποστηρικτικών Πληροφοριών). Συγκεκριμένα, στο Σχ. S3a, παρατηρήθηκαν οι χαρακτηριστικές ζώνες προσρόφησης της simazine που αντιστοιχούν σε NH τέντωμα (3260 cm−1 ) και C N τέντωμα (1637, 1565 και 1406 cm−1 ) [27]. Τα Σχ. S3b και c δείχνουν ότι, μετά την προσρόφηση του ζιζανιοκτόνου, η ζώνη προσρόφησης στα 3440 cm−1, που αντιστοιχεί στην τάνυση –ΟΗ κάθε οξειδίου, μειώθηκε περισσότερο ή λιγότερο έντονα λόγω πιθανών αλληλεπιδράσεων συντονισμού simazine-Fe2O3 και αντιδράσεις οξέος-βάσης ή αντικατάσταση του ζιζανιοκτόνου με μόρια νερού στο σύμπλοκο εξασυντεταγμένων οξέος [Al(H2 O)6 ]3+, αντίστοιχα. Σε μια μελέτη ρόφησης-εκρόφησης της ατραζίνης και της σιμαζίνης με μοντέλα κολλοειδών συστατικών του εδάφους, οι Celis et al. [25] απέδειξε ότι ο φερριϋδρίτης δεν προσροφά ζιζανιοκτόνα τριαζίνης. Η ενισχυμένη απορρόφηση αυτών των ζιζανιοκτόνων στον μοντμοριλλονίτη μετρήθηκε μετά την αύξηση της επιφανειακής οξύτητας του πηλού. Αντίθετα, ένα προϊόν πλούσιο σε άνθρακα (βιοκάρβουνο) που παρήχθη από βιομάζα μέσω πυρόλυσης προσρόφησε μια ποσότητα σιμαζίνης ~2480 ␮mol kg−1 [26].

3.5. Regeneration of Al2 O3 and Fe2 O3 In wastewater treatment that includes the sorption process, the regeneration of the sorbent is essential. Today, the reuse of sorbents in many applications through the regeneration of their adsorption properties is an economic necessity. Desorbents (e.g., sodium hydroxide solution) are commonly used to recover sorbents such as Fe- and Al-based supports [28]. However, the use of a desorption agent has some disadvantages because it increases operating costs, and the waste solution containing NaOH discharged from the regeneration of the sorbent causes environmental pollution. The combustion method can be considered as an alternative way of sorbent regeneration, since it avoids the use of dangerous desorption agents. To evaluate the feasibility of this choice, Al2O3 and Fe2O3 were annealed at 500 ◦ C for 1 hour. Then, in order to determine whether the textural properties are preserved, the physical characterization of heat-treated oxides was performed by analyzing the relative N2 adsorption-desorption isotherms. The authors previously analyzed the porosity of Al2O3 and Fe2O3 [11]. Here we present a comparison with the physical properties of heat-treated samples. The designation used for the samples refers to the chemical formula followed by a number indicating the heat treatment temperature, ie Al2 O3 500 and Fe2 O3 500. samples before heat treatment are simply labeled with the chemical formula as in the text. The perfect agreement between the isotherms obtained with Al2O3 500 and Al2O3 [11] showed that the mesoporous structure was not destroyed by annealing. Adsorption isotherms were prepared using the BET method to obtain the corresponding surface areas listed in Table 3 along with the total pore volume and estimated average pore diameter. As can be clearly seen from the data in Table 3, all textural properties of Al2O3 are well preserved after heat treatment. These results lead us to consider the combustion method as an effective option for alumina regeneration. In contrast, the thermal stability of Fe2O3 was not comparable to that of Al2O3 in terms of pore structure. In fact, the heat treatment strongly changed the textural properties of the sample and a drastic collapse of the surface was observed (Table 3). In order to observe to what extent the iron oxide pore size distribution was modified by the annealing process, the desorption data were processed using the BJH method. A comparison between the pore size distributions of Fe2 O3 and Fe2 O3 500 samples (Figure S4 in the Supporting Information) clearly shows that the mesoporous structure is completely destroyed by heat treatment at 500 ◦ C, rendering the iron oxide irreparably burned.

4. Conclusions In this study, two mesoporous metal oxides, Al2O3 and Fe2O3, showed different adsorption capacities of simazine, a very persistent herbicide. In particular, the optimal pH for sorption was found to be 6.5 for Al2O3 and 3.5 for Fe2O3. The different sorption capacities of the two oxides are explained by considering a set of factors that significantly influence the sorption process, such as surface area and porosity. The adsorption kinetics is described by a pseudo-secondary model, which shows that Fe2O3 adsorbs simazine faster than Al2O3. Finally, we have shown that Al2O3 can be regenerated by incineration and can be considered for industrial treatment systems, for the effective removal of simazine from aquatic environments and finally for pesticide pollution mitigation.

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Additional data The chemical formula of simazine and its analytical determination, Fourier transform infrared spectroscopy (DRIFTS) analysis, Al2O3 and Fe2O3 pore size distribution, and Fe2O3 pore size distribution at 500 ◦ C are reported. Acknowledgments This manuscript is a contribution of DiSSPAPA 248. Appendix A Supplementary data Supplementary data related to this article can be found in the online version at doi:10.1016/j.jhazmat.2011.09.022. References [1] S. Baup, C. Jaffre, D. Wolbert, A. Laplanche, Adsorption of pesticides on granular activated carbon: determination of surface dispersants using simple batch experiments, Adsorption 6 (2000) 219-228. [2] F. Bruna, I. Pavlović, C. Barriga, J. Cornejo, M.A. Ulibarri, Adsorption of the pesticides carbetamide and amitrone on organohydrotalcite, Appl. Clay Sci. 33 (2006) 116-124. [3] G. Crini, The latest development of polysaccharide-based materials used as adsorbents in wastewater treatment, Prog. Polym. Sci. 30 (2005) 38-70. [4] S. Wang, Y. Peng, Natural zeolites as effective adsorbents in water and wastewater treatment, Chem. Meadow. J. 156 (2010) 11–24. [5] N. Ratola, C. Botelho, A. Alves, Using pine bark as a natural adsorbent for persistent organic pollutants-investigation of the adsorption of lindane and heptachlor, J. Chem. Technol. biotechnology. 78 (2003) 347-351. [6] H. Yoshitake, T. Yokoi, T. Tatsumi, Adsorption of chromate and arsenate by amino-functionalized MCM-41 and SBA-1, Chem. Mater. 14 (2002) 4603-4610. [7] C. Lee (ed.), Adsorption Science and Technology, World Scientic, Singapore, 2003, p. 605-609. [8] Y. Kim, C. Kim, I. Choi, S. Rengaraj, J. Yi, Arsenic removal using mesoporous alumina prepared by casting method, Environ. Sci. Technol. 38 (2004) 924-931. [9] M. Anbia, N. Mohammadi, K. Mohammadi, Fast and effective mesoporous adsorbents for the separation of toxic compounds from aqueous media, J. Hazard. Mater. 176 (2010) 965-972. [10] P. Wang, I.M.C. Lo, Synthesis of mesoporous magnetic ␥-Fe2 O3 and its application for the removal of Cr(VI) from polluted water, Vandres. 43 (2009) 3727-3734.

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[11] V. Addorisio, S. Esposito, F. Sannino, Adsorption capacity of mesoporous metal oxides for the removal of MCPA from polluted water, J. Agric. Food Chem. 58 (2010) 5011-5016. [12] G. Celano, D. Smejkalova, R. Spaccini, A. Piccolo, Interactions of three striazins with humic acids of different structure, J. Agric. Food Chem. 56 (2008) 7360-7366. [13] A. Garcia-Valcarcel, J. Tadeo, Effect of soil moisture on the adsorption and degradation of hexazinone and simazine in soil, J. Agric. Food Chem. 47 (1999) 3895-3900. [14] A.S. Gunasekara, J. Troiano, K.S. God, R.S. Tjeerdema, Simazin's chemistry and destiny, Rev. Surroundings. Contam. Toxicol. 189 (2007) 1-23. [15] J. Troiano, D. Weaver, J. Marade, F. Spurlock, M. Pepple, C. Nordmark, D. Bartkowiak, Summary of California well water sampling to detect pesticide residues from nonpoint sources, J. Environ. qual. 30 (2001) 448-459. [16] T. Hayes, P. Case, S. Chui, D. Chung, C. Haeffele, K. Haston, Pesticide mixtures, endocrine disruption, and amphibian declines: underestimating impacts, Environ. A health perspective. 1 (2006) 40-50. [17] C. Flores, V. Morgante, M. Gonzalez, R. Navia, M. Seeger, Adsorption studies of the herbicide simazine on agricultural soils in the Aconcagua Valley, central Chile, Chemosphere 74 (2009) 1544-1549. [18] M. Lucio, P. Schmitt-Kopplin, Modeling the binding of triazine herbicides to humic substances using capillary electrophoresis, Environ. Chem. Easy. 4 (2006) 15-21. [19] F. Rouquerol, J. Rouquerol, K. Singh, Adsorption by Powders and Porous Solids: Methodology and Application of Principles, Academic Press, London, 1999. [20] D.L. Sparks, Environmental Soil Chemistry, Academic Press, London, 1995. [21] M. Chappell, D.A. Laird, M.L. Thompson, H. Li, B.J. Teppen, V. Aggarwal, C.T. Johnston, S.A. Boyd, Effect of smectite hydration and swelling on the sorption behavior of atrazine, Environ. Sci. Technol. 39 (2005) 3150-3156. [22] C. Garau, D. Quinonero, A. Frontiera, P. Ballester, A. Costa, P.M. Deya, mode of double bond of s-triazine to anions and cations, Org. Easy. 5 (2003) 2227-2229. [23] M. Ozacar, I.A. Sengil, Design of a two-stage batch adsorbent for methylene blue removal to reduce contact time, J. Environ. They manage. 80 (2006) 372-379. [24] C.H. Giles, D. Smith, A. Huitson, General treatment and classification of solute adsorption isotherms. I. Theoret., J. Colloid Interface Sci. 47 (1974) 755-765. [25] R. Celis, J. Cornejo, M.C. Hermosin, W.C. Koskinen, Sorption-desorption of atrazine and simazine by colloidal components in a model soil, Soil Sci. Soc. I'm. J. 61 (1997) 436-443. [26] W. Zheng, M. Guo, T. Chow, D.N. Bennett, N. Rajagoplan, Sorptive properties of green waste biochar for two triazine pesticides, J. Hazard. Mater. 181 (2010) 121-126. [27] L.J. Bellamy, The Infrared Spectra of Complex Molecules, vol. I, 3rd edition, Chapman and Hall, New York, 1975. [28] Y. Kim, B. Lee, J. Yi, Preparation of functionalized mesostructured silicon containing magnetite (MSM) for the removal of copper ions in aqueous solutions and its magnetic separation , Sep. Sci. Technol. 38 (2003) 2533-2548.

Journal of Hazardous Materials 196 (2011) 287-294

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Use of arc steel slag as raw material for low energy belite cements R.I. Iacobescu a, D. Koumpouri b, Y. Pontikes c, R. Saban a, G.N. Angelopoulos b,∗ a b c

Department of Materials Science and Engineering, Polytechnic University of Bucharest, Splaiul Independentei 313, 060032 Bucharest, Romanian Laboratory of Materials and Metallurgy, Department of Chemical Engineering, University of Patras, 26500 Rio, Greece Department of Metallurgy and Materialscitolivenekeneering Kasteelpark 44, bus 44 , Arenberg B -3001 Heverlee (Leuven), Belgium

article

i n f o

Article history: Received July 5, 2011 Received in revised form September 7, 2011 Accepted September 7, 2011 Available online September 12, 2011 Keywords: electric arc rust Belite cement

ab s t r a c t This paper investigates the use of electric arc steel slag (EAFS) for the production of low-energy belite cements. Three types of clinker were prepared with 0 wt. % (BC), 5 wt. % (BC5) or 10 wt. % (BC10) EAFS. The design of the raw mixes was based on the composition indices of lime saturation factor (LSF), alumina ratio (AR) and silica ratio (SR). The temperature of the clinker was tested in the range of 1280-1400 ◦ C. Firing was carried out at 1380 ◦ C based on the results on free lime and microstructure development. In order to activate the belite, the clinkers were quickly cooled by blowing air and simultaneously crushed. The results show that the microstructure of the produced clinkers is dominated by belite and halite crystals, where tricalcium aluminate and tetracalcium aluminum ferrite are present as microcrystalline intermediate phases. The produced cements showed low early strength development as expected for compositions rich in belite. However, the 28-day results were 47.5 MPa, 46.6 MPa, and 42.8 MPa for BC, BC5, and BC10, respectively. These values are comparable to OPC CEMI 32.5 N (32.5–52.5 MPa) according to EN 197-1. Fast hardening was also observed, especially in the case of BC10, while the stability did not exceed 1 mm. © 2011 Elsevier B.V. All rights reserved.

1. Introduction In recent years, the cement industry has seen dynamic growth, with most of the activity taking place in developing economies. Despite the economic turmoil, population growth and the resulting need for housing, along with public investment in infrastructure, are strong drivers to offset the decline in the cement market. Worldwide, cement production increased from 2568 Mt in 2006 to 3294 Mt in 2010 [1]. Like any industrial activity, cement production inevitably has its own footprint on the environment. Estimates show that cement production is responsible for 5-7% of global CO2 emissions [2,3]. If all greenhouse gases emitted by anthropogenic activities are included, the cement industry contributes about 3% of total anthropogenic greenhouse gas emissions [2]. This is mainly a result of the fuel used to produce the required energy, estimated at 0.37 kg/kg clinker, and the carbonation of limestone (CaCO3) that takes place during cement production, estimated at 0.53 kg/kg clinker CO2 [2]. The reduction of limestone in raw flour and thus the change in its chemical composition can therefore lead to lower CO2 emissions. This potential has led to increased scientific interest in innovators

∗ Corresponding author. Tel.: +30 2610969530; fax: +30 2610990917. E-mail address:[email protected](G.N. Angelopoulos). 0304-3894/$ – see cover page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.024

types of cement [4], more specifically in belite-rich cements in the last 20 years [5,6]. This type of cement, unlike conventional OPC, contains a higher percentage of belite (C2 S) and a lower percentage of halite (C3 S). To achieve the desired percentage of C2 S and C3 S, the lime saturation factor (LSF) must be between 78% and 83% [7]. The environmental benefits of belite cements compared to OPC can be summarized as follows: energy savings can be increased up to 16% [8], combustion temperatures can be reduced by 6-10%, and emitted CO2 and NOx levels can be reduced [9 , 10] . However, the early strength of such cements is lower and the grinding energy can be increased due to the hardness of C2 S. By combining the production of belite cements with alternative raw materials as a substitute for limestone, such as metallurgical slag, further advantages may be possible. arise. was achieved [11]. Different types of slag are produced during the production of iron and steel. These include blast furnace (BF), basic oxygen furnace (BOF), electric arc furnace (EAF) and stainless steel slag (SS-EAF and SS-AOD). In the world, almost 50 Mt of steel slag is produced annually, and in Europe 12 Mt annually [12]. Of this, about 65% is used in appropriate fields of application, primarily construction, while the rest is stored or used for other micropurposes [13]. About 37% of steel slag produced in Europe in 2010 was used for cement production [14]. Today, more than 40% of global steel production takes place in the EAF [15] and is associated with slag production of 20 Mt/year. Greece has a cement production capacity of about 18 Mt/year [16],

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and steel production capacity of 3.5 Mt/year. Annual production of EAF slag (EAFS) varies from 300,000 t/year to 400,000 t/year. Of the total annual processed amount of EPP approx. 55% for the production of large aggregates for road construction. The main environmental problems associated with EAFS waste disposal are slag "dusting" and leachate discharge. So, apart from the needs of the cement industry, steel producers have their own incentives to find uses for their slag. In principle, there are two methods of incorporation of slag in cement production: either in raw flour or in a later stage, as (latent) hydraulic or pozzolanic material [17]. Previous studies show that the addition of EAFS up to 10 wt% in raw flour is effective without any adverse effects on the technical properties of the resulting cement [17]. Other authors achieved similar results in terms of melting, microstructure as well as hydration properties of the final clinker with the addition of 10.5 wt% EAFS. In addition to the above, other authors suggest that the addition of EAFS in the production of clinker will reduce the sintering temperature of raw flour and the theoretical heat consumption [18]. Additionally, cement containing steel slag such as BOF and/or EAF can also have improved corrosion resistance than conventional Portland cement [19]. Finally, combinations of EAFS, BOF and AOD slag were investigated for the production of sulfo-alumina belite cement with encouraging results [20]. Despite the work in the field, to our knowledge no studies have focused on the use of EAFS for cements rich in belite. Regarding the disadvantages, the content of heavy metals (Cr, V, etc.) in the steel slag is of concern. In the European Union (EU), the Cr(VI) Directive came into force in 2005 and prohibits the use or delivery of cement containing more than 2 ppm of water-soluble chromium in bulk cement [21]. Typically, Cr(VI) compounds are more soluble in water (although insoluble ones also exist), so they are more likely to participate in leaching. Numerous adverse health effects are associated with exposure to Cr(VI), of varying intensity. According to NIOSH [22], all Cr(VI) compounds are considered probable occupational carcinogens. Reducing agents, such as ferrous sulfate, either monohydrate (FeSO4·H2O) or heptahydrate (FeSO4·7H2O), or stannous sulfate (SnSO4) are added to control the oxidation state of chromium [23]. This paper investigates how EAFS can be used as a raw material for the production of low-energy white cements. The produced clinkers were characterized by SEM/EDS and Rietveld QXRD. Water requirement, initial setting time, stability and compressive strength were measured for both cement and cement paste. The hydration behavior of these cements, as well as their washing ability, were investigated in a separate paper.

2. Materials and methods The raw materials for preparing the first meals were limestone, flysch and EPV. Chemical analysis was performed by X-ray fluorescence spectrometry (XRF, Philips PW 2400). The crystalline phases of the starting materials were identified by X-ray diffraction analysis (D5000 Siemens). Qualitative analysis was performed with DIFFRACplus EVA® software (Bruker-AXS) based on the powder diffraction ICDD file. Mineral phases were quantified using the Rietveld quantification routine with TOPAS® software (Bruker-AXS). This routine is based on the calculation of the individual mineral phase pattern and refinement of the pattern using a non-linear least squares routine [24]. Numerous corrections, including adjustments to instrument geometry, background, sample offset, detector type, and mass absorption coefficients of the purified phases, were applied to obtain the best sample fit. Diffraction patterns were

measured over the 2 range of 5–70◦ using CuK␣ radiation at 40 kV and 30 mA, with a step size of 0.01◦ and a step time of 1 deg/min. The design of the raw meals was based on the predictions of the Bogue equations. For the production of high-quality cement, the lime saturation factor (LSF) was adjusted between 78% and 83% [7], while the proportions of alumina (AR) and silica (SR) varied from 1.00% to 1.87% and 1.96 % to 3.29%, similar to those used in OPC production. The quality indices for LSF, AR and SR were calculated according to Eq. (1)–(3) [5,23]:

LSF =

%CaO 2,8 ∗ %SiO2 + 1,2 ∗ %Al2O3 + 0,65 ∗ %Fe2 O3

(1)

SR =

%SiO2 %Al203 + %Fe203

(2)

AR =

%Al203%Fe203

(3)

Based on the above and the chemical analysis of the raw materials, the MS Excel © worksheet was used to derive the composition of raw meals. Three types of clinker were produced: one as reference (named BC), the second with the addition of 5 wt% EALS (BC5) and the third with the addition of 10 wt% EAFS (BC10). The achieved proportion of flour in limestone/flysch/EAFS was in weight percentages: 84.0/16.0/0.0, 80.5/14.5/5.0 and 77.0/13.0/10.0 for BC, BC5 and BC10. The quality index results are shown in Table 3. BC10 shows the maximum in terms of EAFS addition (10 wt%), while LSF remains within the desired limits. The mineralogical phases of the clinker were also calculated using the Rietveld method, with estimates obtained from the Bogue equations (table 4). For the preparation of clinker, the raw materials were ground individually in a Siebtechnik® planetary mill to a particle size below 90 µm. After mixing and homogenization, pellets with a diameter of approx. 15-20 mm formed by hand with minimal addition of water. The pellets were dried for 24 hours at 110 ◦ C, followed by calcination at 1000 ◦ C for 4 hours. Clinker firing was carried out in a Nabertherm® Super Kanthal type resistance furnace at 1380 ◦ C. The optimum clinker temperature was determined by firing tests at 1280 ◦ C, 1300 ◦ C, 1320 ◦ C, 1350 ◦ C, 0, 0 1350 ◦ C ◦ C, with a soaking time of 40 min to determine the free lime content according to ASTM C114-03, as well as SEM observations that assessed the quality of the clinker. In order to stabilize the ␣ - and ␤ - C2 S polymorphic forms, rapid cooling with simultaneous air blowing and hammer crushing was used. Clinkers were characterized by QXRD and SEM/EDS microanalysis (Jeol JSM 6300 and LINK PentaFET 6699, Oxford Instruments). Carbon coated samples were used, cut, polished and etched with 1% Nital. All EDS analyzes were performed far from the phase boundaries. In the case of belite and halite crystals, analyzes were carried out in situ. In the case of the interphase, due to the microcrystalline texture of the individual phases finely distributed within the amorphous layer, the area of approx. 4 ␮m × 4 ␮m from the analyzed boundaries of halite and belite. Standard LINK ISIS templates were used. For the production of cement, clinkers were ground through the aforementioned planetary mill to a fineness in the range of 4000-4100 cm 2 /g. After grinding, 5% of gypsum with a grain size of less than 90 µm was added. The specific surface area (Blaine method) was measured according to EN 196-6 [25], the setting time and stability according to EN 196-3 [26] and the compressive strength according to EN 196-1 [27].

R.I. Iacobescu et al. / Journal of Hazardous Materials 196 (2011) 287–294 Table 1. Chemical composition of raw materials (wt.%). Oxides

EAFS

Limestone

boca

CaO FeOtal SiO2 Al2 O3 MnO MgO Cr2 O3 P2 O5 TiO2 SO3 Cl BaO Na2 O K2 O V2 O5 LOI

32,50 26,30 18,10 13,30 3,94 2,53 1,38 0,48 0,47 0,44 0,14 0,14 0,13 n.d. 0,06 0,00

48,90 1,00 9,00 1,36 at 0,65 n.p. do. do. do. do. do.

5,55 5,90 58,25 13,75 n.p. 2,86 n.a. n.a. n.a. 0,05 n.a.p.a.

In total

99,91

99,16

99,76

LOI, loss on ignition. n.d., not specified.

3. Results and discussion 3.1. Characterization of raw materials XRF chemical analysis of raw materials, limestone, flysch and EAFS are given in Table 1. EAFS was observed to contain elements such as Cr, P, Ti, S and Ba, which are considered additives for the activation of belite. The introduction of such ions into the C2S crystal lattice can stabilize the ␣ - and ␤ - polymorphs. ␣ -C2 S is more active than ␤-C2 S [7,28]. XRD analyzes of raw materials are shown in Figure 1, while the results of semi-quantitative mineralogical analysis are shown in Table 2. The main identified mineralogical phases are calcite and quartz for limestone, quartz, illite, dolomite, albite and clinochlorite pecam. The EAFS contains significant amounts of larnite (␤-belite), galena, wüstite, magnetite and brownmillerite. Rietveld analysis results are 41.0 wt%, 14.7 wt%, 12 wt%, 10.0 wt% and 9.4 wt% for the above phases.

289

Table 2 Mineralogical composition of raw materials, weight, according to Rietveld analysis, normalized. Limestone Limestone Quartz Illite Microcline Muscovite Kaolinite Hematite Clinochlore Cristobalite Total

Fleece 90.5 5.7 1.1 1.1 0.6 0.5 0.2 0.2 0.1

Illit Kvarts Kaolinit Dolomit Albit Kalcit Mikroklin Muskovit Klinoklorit Hæmatit

100

EAFS 34,1 28,2 8,5 6,8 6,7 5,9 5,7 2,2 1,3 0,6

Larnite Gehlenite Wüstite Magnetite Brownmillerite Mayenite Merwinite Spinel

100

41,0 14,7 12,0 10,0 9,4 7,2 3,7 2,0

100

3.2. Clinker quality depending on the firing temperature The composition of the produced clinker as well as the quality indices are shown in table 3. In the firing tests, the maximum proportion of free lime was 1.6% by weight. for both BC and BC10 at 1280 ◦ C. For the temperature range 1300–1400 ◦ C, free lime varied from 0.4 wt% to 0.2 wt% for all tested mixtures. For firing above 1300 ◦ C, the free lime values are therefore well below the commonly defined limit of 1 wt% for OPC clinker. In fig. Figures 2 and 3 show backscattered electron images revealing the development of the clinker microstructure at different firing temperatures for BC and BC10. In both cases, for temperatures up to 1320 ◦ C, the clinker microstructure is poorly developed. A prolonged interphase is also observed. At temperatures above 1350 ◦ C, the effects of dissolution and transport through the melt are enhanced: the precipitation of stable, rounded belite is visible in association with the formation, to a lesser extent, of angular halite. The resulting microstructures are characterized by uniform phase distribution and growth. Firing at 1400 ◦ C has no visible difference in morphology and phase growth compared to 1380 ◦ C. Comparing the microstructures of BC and BC10, the addition of slag does not promote the formation of halite and promotes the formation of an intermediate phase as well as Table 3 The composition of the raw metal, chemical composition of the produced clinker was obtained and results of quality indicators.

Sirovina EAFS Vapnenac Laneni oksidi SiO2 Al2 O3 Fe2 O3 CaO MgO K2 O Na2 O SO3 MnO Cr2 O3 P2 O5 TiO2 Cl BaO V2 O5 Sl. 1. XRD uzorici sirovina. Glavni identificirani minerali su: 1, kalcit (CaCO3). 2, kvarc (SiO2 ); 3, ilit ((K,H3O)(Al,Mg,Fe)2(Si,Al)4O10((OH)2,(H2O))); dolomite (CaMg(CO3)2); 5, albit (NaAlSi3O8); 6, klinoklorid 4, (Mg2,5Fe1,65Al1,5Si2,2Al1,8010(OH)8); 7, larnit (␤-Ca2SiO4); 8, galenite (Ca2Al(AlSi)O7); 9, wustit (FeO); 10, magnetite (Fe3O4); 11, smeđi milerit (Ca2(AlFe3)2O5); 12, maenit (Ca12Al14O33).

Indicators of total quality of LSF AR SR

0 wt.% BC

5 wt.% BC5

10% BC10

0,0 84,0 16,0

5,0 80,5 14,5

10,0 77,0 13,0

25,38 5,03 2,68 63,09 1,51 0,79 0,39 0,01 0,00 0,00 0,00 0,00 0,00 0,00 0,00

24,41 5,52 4,38 61,47 1,57 0,71 0,36 0,04 0,29 0,10 0,04 0,03 0,01 0,01 0,00

23,48 6,00 6,00 59,92 1,62 0,63 0,34 0,07 0,57 0,20 0,07 0,07 0,02 0,02 0,01

98,88

98,95

99,01

80,13 1,87 3,29

79,00 1,26 2,47

78,10 1,00 1,96

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Fig. 2. Backscattered images of preliminary recordings. BC clinker produced at different firing temperatures: (a) 1280 ◦ C, (b) 1300 ◦ C, (c) 1320 ◦ C, (d) 1350 ◦ C, (e) 1380 ◦ C and (f) 1400 C

belite The crystal size varies from 5 µm to 50 µm. These values are typical for good quality clinker. According to the above results, it was decided to fire the clinker at 1380 ◦ C. 3.3. Characterization of clinker XRD patterns of produced clinker are shown in Fig. 4. Table 4 shows mineralogical compositions calculated by Rietveld and estimates obtained from Bogue's equations. As expected, there are differences between the results obtained by Bogue and Rietveld, the most significant being Bogue's underestimation of C3 S and overestimation of C2 S. This is attributed, among other things, to

in the fact that Bogue's method is based on ideal stoichiometries of clinker phases without consideration of solid solutions, and it also implies a specific course of reaction and solidification. On the other hand, Rietveld analysis can reflect changing conditions caused by different raw materials as well as non-equilibrium conditions during firing and cooling. Nevertheless, Bogue correctly predicted the qualitative trend. Based on the X-ray diffraction results, the main mineralogical phases were identified: halite (C3 S), belite (C2 S), tricalcium aluminate (C3 A) and tetracalcium aluminum ferrite (C4 AF). This particular polymorphism is important because it affects hydration and consequently the development of microstructure and mechanical properties. The primary polymorphic form of stabilized halite is tri-

Fig. 3. Backscattered images of preliminary recordings. BC10 clinker produced at different firing temperatures: (a) 1280 ◦ C, (b) 1300 ◦ C, (c) 1320 ◦ C, (d) 1350 ◦ C, (e) 1380 ◦ C and (f) 140 ◦ C .

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291

Fig. 4. X-ray samples of prepared tiles. Main minerals identified: 1, C3 S; 2, C2S; 3, C3A; 4, C4 AF.

clinical (T), which decreases as the slag content increases towards monoclinic (M). This is expected to influence early strength development towards lower values [29]. Rhombohedral (R) alite reacts somewhat faster compared to T and M polymorphic forms [29,30], but without a significant effect in this case due to its low content. As for belite, mainly polymorph ␣ was detected. ␥- and ␤-C2 S polymorphs were also detected and increased with increasing slag content. Beta, ␤- and ␣ -C2 S play an important role in the last days of hydration. In contrast, ␥C2 S does not show significant hydraulic properties. C3 A was formed as cubic and orthorhombic in BC, while only orthorhombic was found in BC5 and BC10. The predominance of orthorhombic C3 A in the last two clinkers is probably due to the higher sulfate content in EAFS, which inhibits the formation of cubic C3 A [31]. The identified orthorhombic polymorph C3 A will react faster in the presence of gypsum than its cubic counterpart [32]. In general, an increase in EAFS content leads to a decrease in C3 S and C3 A and an increase in C4 AF. this is attributed to the high iron content of the slag.

Table 4 Estimated (Bogue) and calculated (Rietveld) mineralogical composition of the prepared clinkers. Phases

npr. Rietveld

BC5 Bogue

Rietveld

BC 10 Bog

Mose

1,0 0,3 33,3 0,7

Ukupno C2 S-␣’ C2 S-␤ C2 S-␥

35,30 45,5 1,9 0,0

27,43

28,90 41,6 4,0 1,5

22.39

21,80 42,2 5,8 0,8

17,59

Ukupno C3 A – Kubik C3 A – Red.

47,40 2,7 6,0

54,96

47,10 0,0 4,9

56,16

48,80 0,0 4,3

57,32

8,70 7,50 0,4 0,7 100

9,13 8,49 0,0 0,0 100

4,90 18,50 0,0 0,6 100

7,54 13,90 0,0 0,0 100

4,30 24,10 0,0 1,0 100

6,01 19,08 0,0 0,0 100

Total C4 Lime MgO Total

3,4 1,3 20,8 3,4

Rietveld

C3 S (M1) C3 S (M3) C3 S-T C3 S-R

4,0 1,5 15,4 0,9

Fig. 5. Backscattered images of clinker in polished section: (a) BC, (b) BC5 and (c) BC10.

Backscattered electron images (BEI) are shown in Fig. 5a, b and c for BC, BC5 and BC10 respectively. In all cases, the microstructures consist mainly of well-developed type I belite crystals according to Insley's classification [33]. Their diameter varies from 5 m to 50 m. Most belite crystals show complex double laminae. The streaks are formed as a result of phase transformation during cooling. In type I belite, it has been reported that it is a skeletal structure, not a polysynthetic twin, consisting of beta and alpha forms of belite [33]. Crystals with parallel stripes are less noticeable. There are also angular, euhedral and subhedral halite crystals. In the case of BC10, halite crystals

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there are inversions during cooling that cause twinning, resulting in strain accumulation [34] and (b) they are caused by etching [33]. Representative EDS microanalyses of C3S and C2S crystals and the interphase for each clinker produced are shown in Table 5. The ferrite and alumina phases are not shown separately due to difficulties arising from their microcrystalline texture, but the interphase analysis is. marked as C3 A + C4 AF in the table. As a general observation, halite and belite are observed to become Fe-enriched as the slag content increases, while Ba, Cr, Ti and P are more likely to be found in belite crystals. According to the available QXRD and SEM/EDS results and under the adopted experimental conditions, no clear conclusions can be drawn about the effect of slag addition on the vellite crystals, although it is about the richness of elements such as P, S, Cr. known as belite stabilizers. But that will be the subject of the upcoming announcement about the composition and substitutions in certain phases. 3.4. Measurements of specific surface area, setting time and stability Table 6 shows the results of water demand, initial setting time and stability of cement slurries. For the production of cement paste of standard composition [26], the need for water was 27.6% by weight for all cases. The initial setting times obtained for BC, BC5 and BC10 were 240 min, 170 min and 20 min according to EN 197-1 in the first 2 cases. It has been observed that the use of slag shortens the setting time, since BC10 behaves as a fast setting cement. This is attributed to an increase in the molten phase, which is formed by cooling larger amounts of C4 AF, and to the interaction with C3A and gypsum during hydration. Specifically, since the early hydration of the cement is mainly controlled by the amount and activity of C3 A, the separation is balanced by the amount and type of sulfate intermediate with the cement. Tetracalcium aluminum ferrite (C4 AF) reacts like C3A, i.e. forms ettringite in the presence of gypsum. Higher amounts of C4 AF in slag clinkers (more than twofold and threefold increase for BC5 and BC10, respectively, compared to BC) also consume higher amounts of gypsum. Therefore, it is likely that coagulation occurs due to the uncontrolled reaction of C3 A after the sulfate has been depleted by reaction with C4 AF. The addition of larger amounts of gypsum or retarders, preferably of an organic nature, could control the setting behavior. Expansion was 1 mm for all prepared cements. To obtain Blaine of 4000 cm2/g, 4080 cm2/g and 4057 cm2/g, the required grinding times were 60 seconds, 80 seconds. and 103 sec. for BC, BC5 and BC10 respectively. In particular, increased slag addition results in increased grinding time for comparable fineness. 3.5. Compressive strength of cements BC, BC5 and BC10 Fig. 6. Backscattered images of clinker on the fractured surface: (a) BC, (b) BC5, and (c) BC10.

sometimes they contain inclusions of belite. The crystallized under-cooled interphase shows a microcrystalline texture and consists mainly of a mixture of ferrite and C3 A. It partially separates the primary crystals of halite and belite. The addition of slag favors the formation of belite and ferrite phases, while it favors the formation of halite. Ferrite crystal tubes exposed on the fractured surface are shown in Fig. 6b and c for BC5 and BC10, respectively. The voids in their structure are probably occupied by aluminum. The structure of lamellar velite is easily visible. In Figure 6b, microcracks developed in belite crystals are clearly visible. Reasonable hypotheses for the formation of these microcracks are: (a) since dicalcium silicate crystals are considered complex, they contain point defects;

Fig. Fig. 7 shows the compressive strength results. As expected for white cements, the development of strength in the first days is significantly lower than with OPC. For BC, the two-day results were 6.5 MPa, while for BC5 and BC10 they were even lower at 2.5 MPa and 1.6 MPa, respectively. However, the 28-day results for BC, BC5, and BC10 were 47.5 MPa, 46.6 MPa, and 42.8 MPa, respectively, which are comparable to OPC CEMI 32.5N (32.5–52.5 MPa) according to EN 197-1 [35] . These results are consistent with previously published results [36-39]. The low early strength development observed in cements with slag addition is attributed, as in the case of rapid hardening, to extensive ferrite formation. The higher results for BC in the first days are attributed to the higher content of C3 S and C3 A compared to BC5 and BC10. Compared to other belite cements that contain waste, belite cement with EAFS has

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293

Table 5 Typical EDS microanalyses for clinker: (a) BC, (b) BC5 and (c) BC10, by weight. Phase

f.Kr

5 f.Kr

10 BC

C3 S

C2 S

C4 AF + C3 A

C3 S

C2 S

C4 AF + C3 A

C3 S

C2 S

C4 AF + C3 A

NaO MgO Al2 O3 SiO2 P2 O5 SO3 K2 O CaO TiO2 MnO FeOtal V2 O5 BaO Cr2 O3

0,10 0,40 1,23 24,05 0,00 0,00 1,43 71,25 0,00 0,00 1,54 0,00 0,00 0,00

0,20 0,36 1,73 30,65 0,00 0,20 1,44 64,10 0,00 0,00 1,32 0,00 0,00 0,00

0,67 1,45 20,74 5,35 0,00 0,08 0,50 59,14 0,00 0,00 12,07 0,00 0,00 0,00

0,11 0,30 1,66 23,66 0,00 0,00 1,36 69,97 0,04 0,36 2,19 0,00 0,00 0,35

0,01 0,36 0,84 30,88 0,23 0,03 1,35 61,38 0,10 0,39 3,22 0,00 0,09 0,22

0,28 1,53 20,04 4,16 0,00 0,14 0,50 56,41 0,06 1,35 15,43 0,00 0,05 0,05

0,09 0,81 2,25 22,49 0,01 0,10 1,50 68,01 0,08 0,94 3,34 0,00 0,03 0,35

0,12 0,97 2,82 30,25 0,43 0,00 0,81 59,03 0,04 0,38 4,52 0,00 0,13 0,50

0,19 1,35 18,99 4,78 0,00 0,13 0,54 53,59 0,09 1,38 18,86 0,02 0,06 0,02

In total

100

• The first compressive strength results are low as expected for white cements, but the 28-day results for BC, BC5 and BC10 were 47.5 MPa, 46.6 MPa and 42.8 MPa respectively, which is comparable to EN 197 -1, OPC CEMI 32.5N. • The addition of slag did not affect the water requirements of the cement and the stability did not exceed 1 mm, although the setting time was reduced for BC10, which behaved like a "flash-set" cement. • The rapid setting observed for slag-added cements is attributed to extensive formation of a ferrite-forming molten phase during cooling and interaction with C3A and gypsum during hydration. Thank you

Fig. 7. Results of compressive strength of BC, BC5 and BC10 cements.

Table 6. Physical properties of cement and cement pastes. Types of cement Specific surface area (cm2 /g) Initial setting time (min) Water requirement (wt%) Hardness (mm)

f.eks. 4000 240 27,6 1

5 f.Kr

10 BC

4080 170 27,6 1

4057 20 27,6 1

R.I. Iacobescu and R. Saban acknowledge the Sectoral Operational Program for the Development of Human Resources 2007-2013. Ministry of Labour, Family and Social Protection of Romania through financial agreement POSDRU/6/1.5/S/16. D. Koumbouris and G.N. Angelopoulos thanks the support of the University of Patras through the "Karatheodoris" research program in 2011. Mr. Pontikes thanks the Research Foundation - Flanders for a postdoctoral scholarship. TITAN Cement Company S.A. and metallurgical industry SOVEL S.A. are grateful for the provision of raw materials as well as their technical assistance. bibliographical references

lower early and late compressive strength than belite red mud of the Bayer process [38] (5.0 MPa 1 day and 53.7 MPa after 28 days) and higher early and late compressive strength than boron waste belite cement [39] (1.5 MPa 1 day and 32.3 MPa after 28 days). 4. Conclusion Production of belite cements with EAFS is feasible and can provide significant environmental benefits. Concretely, the properties of cements are as follows: • Clinkers mostly contain well-formed belite crystals. Halite crystals are also present. • The intermediate phase is a mixture of C4 AF and C3 A and partially separates the primary crystals of halite and belite. • The addition of slag favors the formation of belite and ferrite phases and discourages the formation of halite, as predicted by Bogue in accordance with the requirement of comparable quality indices.

[1] International Cement Review, The Global Cement Report, ninth edition, World Overview, 2011. [2] J.S. Damtoft, J. Lukasik, D. Herfort, D. Sorrentino, E.M. Gartner, Initiatives for Sustainable Development and Climate Change, Cem. Inc. Res. 38 (2008) 115-127. [3] V.M. Malhotra, Global warming and the role of cement admixtures and superplasticizers in reducing greenhouse gas emissions from portland cement production, Int. J. Struc. Meadow. 1 (2010) 116-130. [4] M. Schneider, M. Romer, M. Tschudin, H. Bolio, Sustainable cement production – present and future, Cem. Inc. Res. 41 (2011) 642-650. [5] J.I. Bhatty, F.M. Miller, S.H. Kosmatka, Innovations in Portland Cement Manufacturing, Portland Cement Association, 2004. [6] C.D. Popescu, M. Muntean, J.H. Sharp, Industrial trial production of low-energy belite cement, Cem. Inc. Complex. 25 (2003) 689-693. [7] C.D. Lawrence, Production of low-energy cements, in: Lea's Chemistry of Cement and Concrete, 4th edition, Butterworth-Heinemann, Oxford, 2003. [8] H. Uchikawa, Management strategy in cement technology for the next century. Part 3, World Cem. November (1994) 47. [9] E. Gartner, Industrially interesting approaches to "low CO2" cements, Cem. Inc. Res. 34 (2004) 1489-1498. [10] K. Quillin, Performance of belite-sulfoaluminate cements, Cem. Inc. Res. 31 (2001) 1341-1349. [11] US Environmental Protection Agency, Materials Characterization Document, Supporting Final Rulemaking: Identification of Non-Hazardous Secondary

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[19]

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[23] [24]

[25] [26]

R.I. Iacobescu et al. / Journal of Hazardous Materials 196 (2011) 287-294 materials that are solid waste steel slag (used as an ingredient in the production of clinker and bituminous concrete), in: Steel Furnace Slag, 2011, p. 1-5. I. Akln Altun, Y. Ismail, Investigation of High MgO Steel Furnace Slag as an Additive in Portland Cement, Cem. Inc. Res. 32 (2002) 1247-1249. H. Motz, J. Geiseler, Steel slag products an opportunity to save natural resources, Waste Management. 21 (2001) 285-293. European Slag Association, www.euroslag.org, 2011. International Iron and Steel Institute, IISI Steel Statistical Yearbook, Brussels, Belgium, 2004. http://www.worldsteel.org/. Hellenic Cement Industry Association, http://www.hcia.gr. S.R. Rao, Waste Management Series 7, in: Resource Recovery and Recycling from Metallurgical Wastes, 2006. J. Paceagiu, E. Radulescu, A.M. Dragomir, R. Hotnog, Implications of using steel slag for clinker production: results of laboratory tests, Romanian J. Mater. 40 (2010) 306-314. C. Shi, Corrosion-resistant cement made from steelmaking by-products, Proceedings of the International Symposium on the Utilization of Metallurgical Slag, China Society for Metals, Beijing, 1999, p. 171-178. D. Adolfsson, N. Menad, E. Viggh, B. Björkman, Steelmaking slag as raw material for sulfoaluminate belite cement, Adv. Cem. Res. 19 (2007) 147-156. European Parliament, Directive 2003/53/EC of the European Parliament and of the Council, 2003 US Department of Health and Human Services, Centers for Disease Control and Prevention, National Institute for Occupational Safety and Health, NIOSH, Updated Criteria Document, Occupational Exposure to Hexovalent Chromium, External Review Draft, 2008. B.W. Nicholas, Understanding Cement, WHD Microanalysis Consultant Ltd., UK, 2010. C.R. Ward, J.C. Taylor, C.E. Matuliš, L.S. Dale, Quantification of mineral matter in premium argon coal using Rietveld interactive X-ray diffraction, Int. J. Coal Geol. 46 (2001) 67-82. European Committee for Standardization, EN 196-6, Test methods for cement. Determination of fineness, 1989. European Committee for Standardization, EN 196-3, Test methods for cement. Part 3. Determination of deposition time and sharpness, 1994.

[27] European Committee for Standardization, EN 196-1, Test methods for cement. Part 1. Determination of strength, 1994 [28] S.N. Ghosh, R.B. Rao, A.K. Paul, K. Raina, review. Chemistry of dicalcium silicate minerals, J. Mater. Sci. 14 (1979) 1554-1566. ´ Effect of halite polymorphism on strength [29] T. Stanek, P. Sulovsky, Portland cement, Cem. Inc. Res. 32 (2002) 1169-1175. [30] R.T.H. Aldous, Hydraulic behavior of rhombohedral halite, Cem. Inc. Res. 13 (1983) 89-96. [31] L. Gobbo, L. Sant' Agostino, L. Garcez, C3 A polymorphs associated with industrial clinker alkali content, Cem. Inc. Res. 34 (2004) 657-664. [32] A. Kirchheim, V. Fernàndez-Altable, P. Monteiro, D. Dal Molin, I. Casanova, Analysis of cubic and orthorhombic C3 A hydration in the presence of gypsum and lime, J. Mater. Sci. 44 (2009) 2038-2045. [33] D.H. Campbell, Microscopical Examination and Interpretation of Portland Cement and Clinker, Portland Cement Association, Skokie, IL 60077-1083, USA, 1999. [34] S.N. Ghosh, Advances in Cement Technology: Critical Reviews and Case Studies on Production, Quality Control, Optimization and Use, New Delhi, India, 1983. [35] European Committee for Standardization, EN 197-1, Cement. Part 1. Composition, specifications and conformity criteria for joint cement, 2000. [36] J. Stark, A. Müller, R. Schrader, K.R. Rumpler, Existence Conditions of Hydraulically Active Belite, Zement-Kalk-Gips, Bauverlag GmbH, Wiesebaden, Germany, 1981. [37] I. Vangelatos, Valorization of Red Rud in the Cement Industry, Department of Chemical Engineering, University of Patras, Patras , 2008. [38] I. Vagelatos, Y. Pontikis, G.N. Angelopoulos, Ferroaluminium as a raw material for the production of belite-type "green" cement, in: SERES' 09. I. International Congress on Ceramics, Glass, Porcelain Enamel, Glaze and Pigment, Eskisehir, Turkey, 2009. [39] T Kavas, I. Vagelatos, S. Koyas, Y. Tabak, G.N. Angelopoulos, Alumina and boron wastes as raw materials for belite cement, in: J. Heinrich, C. Aneziris (eds.), 10th Conference and Exhibition of the European Ceramic Society, Berlin, Germany, 2007, p. 1799-1803. .

Journal of Hazardous Materials 196 (2011) 248-254

Content is available on SciVerse ScienceDirect

Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Effect of preparation method, nitrogen source and post-treatment on photocatalytic activity and stability of N-doped TiO2 nanopowder Shaozheng Hu ∗, Fayun Li, Zhiping Fan Institute of Eco-Environmental Sciences, Liaoning Shihua University, Fushun 1130 China

article

i n f o

Article history: Received May 19, 2011 Received in revised form September 6, 2011 Accepted September 6, 2011 Available online September 10, 2011 Keywords: Network nitrogen TiO2 Photocatalysis Stability After processing

a b s t r a c t NH3 plasma, N2 plasma and NH3 flow annealing were used to prepare N-doped TiO2. XRD, UV-vis spectroscopy, N2 adsorption, FT-IR, Zeta potential measurement and XP spectra were used to characterize the prepared TiO2 samples. The nitriding process did not change the phase composition and particle size of the TiO2 samples, but it broadened its absorption edges in the visible light region. Photocatalytic activities were tested in the decomposition of an aqueous solution of a reactive dye, methylene blue, under visible light. Photocatalytic activity and stability of TiO2 prepared by NH3 plasma was much higher than samples prepared by other nitration procedures. The visible light activity of the prepared N-doped TiO2 was improved by increasing the lattice nitrogen content and decreasing the adsorbed NH3 on the catalyst surface. The stability of nitrogen in the lattice of N-doped TiO2 samples was improved after rinsing with HCl solution. A possible mechanism of photocatalysis is proposed. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Nanocrystalline TiO2 has great potential for many applications such as photocatalysis, solar energy conversion and gas sensing [1,2]. However, with a wide bandgap energy of 3.0-3.2 eV, TiO2 cannot be activated to generate photoexcited electrons and holes to drive the redox reaction unless irradiated with UV light. This precludes the use of TiO2 as a photocatalyst that responds to sunlight or even indoor light. Therefore, it is highly desirable to move the absorption edge of TiO2 into the region of visible light. In 2001, Asahi et al. [3] prepared nitrogen-doped TiO2 films by sputtering TiO2 in N2/Ar gas mixture and concluded that doped N atoms reduce the band gap of TiO2 by mixing N 2p and O 2p states, showing activity for acetone and methylene blue decomposition. Since then, N-doping has become a hot topic and is widely researched. Heating TiO2 powder in N2 and/or NH3 at high temperatures is a conventional method for producing nitrogen-doped TiO2 [3]. Besides the energy loss, processing at such high temperatures usually results in a low surface area due to grain growth, which would reduce the number of photoactive sites. Therefore, new strategies for producing nitrogen-doped TiO2 have emerged, such as sputtering [4], sol-gel [5], ion implantation [6], pulsed laser deposition [7], hydrothermal synthesis [8], and plasma processing [9]. ]. proposed recently.

Non-thermal plasma consists of atoms, ions and electrons, which are much more reactive than their molecular precursors. Plasma can cause many reactions that take place effectively only at high temperatures and high pressures, under mild conditions. So far, the literature on the preparation of N-doped TiO2 by plasma treatment has been published [9-13]. Yamada et al. [9-11] investigated the photocatalytic activity of TiO2 thin films prepared by plasma treatment with N2 as nitrogen source. They suggested that the N-substituent doping contributed to the gap reduction and thus absorbed visible light and showed photocatalytic activity. Abe et al. [12] prepared N-doped TiO2 with NH3 (10%)/Ar plasma. The influence of NH3/Ar gas pressure (50, 300 and 1000 Pa) on the physical and photocatalytic properties of the powder was investigated. Miao et al. [13] reported the structural and synthetic properties of TiO2 thin films prepared by N2-H2 plasma treatment. HRTEM results showed that the primitive lattice cells in anatase TiO2 films were deformed after plasma treatment compared to bulk TiO2, which confirmed the N-doping of N2-H2 plasma. From the above literature, it appears that N2 and NH3 were commonly used as nitrogen source to produce N-doped TiO2 during plasma processing. However, little literature has been published comparing N2 and NH3 plasma treated TiO2. In this work, NH3 plasma, N2 plasma and NH3 flow annealing were used to prepare N-doped TiO2. The structural and optical properties of the prepared TiO2 doped with N were compared. The photocatalytic performance was evaluated in the degradation of methylene blue under visible light. A possible mechanism of photocatalysis is proposed.

S. Hu i sur. / Journal of Hazardous Materials 196 (2011) 248–254

249

2. Experimental 2.1. Preparation and characterization

2.2. Photocatalytic reaction Methylene blue (MB) was chosen as a model compound to evaluate the photocatalytic efficiency of the prepared TiO2 particles in aqueous solution under visible light irradiation. A 0.1 g TiO2 powder was dispersed in 100 mL aqueous MB solution (initial CO concentration = 50 ppm, pH 6.8) in an ultrasonic generator for 10 minutes. The suspension was transferred to an improvised glass reactor and stirred for 30 min in the dark to achieve adsorption equilibrium. The concentration of MB at that time was considered the equilibrium concentration of CO absorption. The adsorption capacity of the catalyst on MB is defined by the amount of adsorption of MB on the photocatalyst (CO − CO). In the photoreaction under visible light irradiation, the suspension was exposed to a 110 W high-pressure sodium lamp with a main emission in the range 400-800 nm, and air was blown at a rate of 130 ml min-1 through the solution. Part of the UV light from the sodium lamp was filtered with a 0.5 M NaNO2 solution [14]. The light intensity is 130 mW cm−2. All experiments were performed at room pressure and 30 ◦ C. At given time intervals, 4 mL of the suspension was withdrawn and immediately centrifuged to separate the liquid samples from the solid catalyst. MB concentrations before and after the reaction were measured with a UV-vis spectrophotometer at a wavelength of 665 nm. It is a linear relationship between the absorbance and the concentration of the liquid sample in the experimental concentration range. Therefore, the degradation rate D% is determined as follows: D% =

A0 − A × 100 % A0

TO-PN2

TO-CNH3

P25

2./degree. Fig. 1. XRD patterns of P25 and prepared N-doped TiO2 samples.

where A0 and A are the absorbances of the liquid sample before and after decomposition. 3. Results and discussion It has been noted that the phase composition and particle size of TiO2 have a significant effect on its photocatalytic activity [2]. XRD patterns of P25 and prepared N-doped samples (Figure 1) show that all TiO2 samples are a mixture of anatase and rutile phases. The phase content and particle size of the samples were calculated by their XRD patterns according to the method of Spurr [15] and the Debye-Scherrer equation [16]. The results (Table 1) show that there were no noticeable changes in the phase composition and particle size. To date, the mechanism of N-doping enhancement is still controversial. Asahi et al. [3] concluded that doped N atoms reduce the band gap of TiO2 by mixing N 2p and O 2p states, which proved the activity. Irie et al. [17] argued that an isolated narrow band above the valence band is responsible for the visible light response. Lee et al. [18] suggested that substitutional nitrogen doping would reduce the band gap by coupling O 2p and N 2p orbitals, while interstitial nitrogen doping would create an isolated defect state between the conduction band and the valence band. Fig. Fig. 2 shows the UV spectra of P25 and prepared TiO2 samples doped with N. Compared to the spectra for P25, distinct red shifts of the absorption bands were observed for the previously prepared N-doped TiO2. 1.0

TO-PNH3

0,8

TO-CNH3 TO-PN2

0,6

Abs

TiO2 doping was carried out in a dielectric barrier discharge (DBD) reactor, which consists of a quartz tube and two electrodes. The high voltage electrode was a stainless steel rod (2.5 mm) mounted on the shaft of a quartz tube and connected to a high voltage source. The ground electrode was an aluminum foil wrapped around a quartz tube. For each test, 0.4 g of commercial TiO2 powder (P25) was loaded into a quartz tube. At a constant flow of NH3 (40 mL min−1), a high voltage of 9–11 kV drove the plasma generator with a total input power of 50 V × 0.4 A. The discharge frequency was set to 10 kHz, and the discharge was maintained for 15 min. . After reading, the reactor was cooled to room temperature. The resulting TiO2 sample is designated as TO-PNH3. When N2 was used as a substitute for NH3 following the same procedure in the preparation of TO-PNH3, the product was designated as TO-PN2. For comparison, P25 was burned under a flow of NH3 (40 ml min-1) for 15 minutes at 500 ◦ C. The obtained sample was designated as TO–CNH3. XRD patterns of the prepared TiO2 samples were recorded on a ˚ Rigaku D/max-2400 instrument using Cu K␣ radiation (= 1.54 A). UV-vis spectroscopy measurement was performed on a Jasco V-550 spectrophotometer using BaSO 4 as a reference sample. FT-IR spectra were obtained on a Nicolet 20DXB FT-IR spectrometer in the range 400-2300 cm-1. Zeta potential of the catalyst was measured at room temperature on a Zetasizer Nano S90 (Malvern Instruments). The pH was adjusted by dropwise addition of dilute HCl or NaOH. Photoluminescence (PL) spectra were measured at room temperature with a fluorospectrophotometer (FP-6300) using a Xe lamp as the excitation source. XPS measurements were performed on a Thermo Escalab 250 XPS system with Al K ␣ radiation as the excitation source. Binding energies were calibrated against the C1s peak (284.6 eV) to minimize the effect of sample charge.

Intensity / a.u.

TO-PNH3

P25 0,4

0,2

0,0 300

400

500

600

700

Wavelength / nm (1) Fig. 2. UV spectra of prepared P25 and N-doped TiO2 samples.

800

250

S. Hu i sur. / Journal of Hazardous Materials 196 (2011) 248–254

Table 1. Overview of physical properties of P25 and prepared N-doped TiO2 samples. Sample

Size (nm)

HA (%)a

SBET (m2 g−1 )

Pore volume (cm3 g−1 )

Central pore size (nm)

F.eks. (eV)

Nfrisk (in.%)b

Female (in.%) c

P25 TO–PN2 TO–CNH3 TO–PNH3

28,2 28,5 29,3 28,1

74,6 74,7 74,4 75,1

43 40 36 41

0,07 0,06 0,05 0,06

3,6 3,2 3,1 3,4

3,10 2,92 2,75 2,67

0 1,32 1,64 1,95

0 0,76 1,17 1,91

does

XA represents the phase composition of anatase. Nfresh represents the nitrogen content of the lattice before the photocatalytic reaction. Nused represents the nitrogen content of the lattice after the photocatalytic reaction.

Band gap energies of TiO2 samples calculated according to the method of Oregan and Gratzel [19] (Table 1) show that the prepared TiO2 samples doped with N showed very limited band gap energies. In agreement with the previous result [3,18], this showed that there is a substitute for nitrogen doping in the prepared samples of nitrogen-doped TiO2. It was shown that the bandgap energy decreased in the order: TO-PN2 > TO-CNH3 > TO-PNH3, which is probably a consequence of the different doping content in the prepared N-doped TiO2 samples. Xu et al. [20] prepared N-doped TiO2 by pulsed laser deposition and suggested that a higher red shift of the absorption edge indicates a higher nitrogen concentration. In addition, clear differences in visible light absorption were observed between calcined and plasma treated samples. In the spectrum of TO-CNH3, the apparent absorption is observed at 400-550 nm, which is the typical absorption region for N-doped TiO2 materials. This typical absorption is a consequence of the electronic transition from the isolated N 2p level, which is formed by the incorporation of nitrogen atoms into the TiO2 lattice, into the conduction band [21]. However, the spectra of plasma-treated samples are of course different. Broad absorptions in the visible light region were observed in the spectra of TO-PN2 and TO-PNH3. Huang et al. [22] prepared visible light-responsive TiO2 by nitrogen plasma surface treatment and found a corresponding broad absorption in the visible light region. Abe et al. [12] prepared N-doped TiO2 using NH3/Ar plasma and suggested that such broad absorption is attributed to the presence of Ti3+ which can be formed by plasma treatment. It was observed that TO-PNH3 showed a much stronger broad absorption in the visible light region than TO-PN2. This is probably because NH3 plasma consists not only of various forms of active nitrogen, but also of excited hydrogen, so that Ti4+ is easily reduced, thereby generating more Ti3+. Therefore, according to the conclusion of Lee et al. [18], it was suggested that the N-substituent and intermediate Ndoping were present simultaneously in TO-CNH3, causing narrowing of the band gap and significant absorption at 400-550 nm, while only N-doping substitution was present in TO-PN2 and TO -PNH3. That

N 1s

TO-PN2

Intensity / a.u.

a b

TO-PNH3

TO-CNH3

396

398

400

Fig. 3. XP spectra of the prepared N-doped TiO2 samples in the N1s region.

The broad absorption of TO-PN2 and TO-PNH3 in the visible light region was due to the presence of Ti3+ caused by N-doping. XPS is an effective surface investigation technique for the characterization of elemental composition and chemical states. According to previous literature [10,11], the peaks around 396 and 400 eV are attributed to the formation of lattice nitrogen and other surface N species such as N-N and N-O bonds. XP spectra in the N1s region (Figure 3) showed that most of the N species in the prepared TiO2 using NH3 as a nitrogen source exist in lattice nitrogen, while other surface N species such as N-N and N-O bonds dominate in TO-PN2 using N2 as a source nitrogen. The lattice nitrogen content calculated from the XPS data is shown in Table 1. The Nfresh content decreased in the order: TO–PNH3 > TO–CNH3 > TO–PN2, which showed that NH3 plasma treatment was more effective than the other two methods

(I)

(U)

From 1 d

Intensity / a.u.

From

TO-PNH3 DO-CNH3

TO-PNH3 TO-PN2 TO-CNH3

TO-PN2

P25

456

458

460

462

Binding energy / eV

464

466

402

Binding energy / eV

529

530

531

532

Binding energy / eV

Fig. 4. XP spectra of P25 and prepared N-doped TiO2 samples in the Ti 2p (A) and O1s (B) region.

533

534

S. Hu i sur. / Journal of Hazardous Materials 196 (2011) 248–254

251

P25 TIL-CNH3

TO-P HN3

TO-PN2

1070 cm

TO-C HN3

-1

1224 cm

-1

P25 1402 cm-1 1620 cm-1

655 cm-1

Zeta potentials / mV

Fravær / a.u.

TO-PNH3 0 2

-10

-20

-30

pH 400

600

800

1000

1200

1400

1600

Wave number / cm

1800

in 2000

2200

-1

Fig. 6. Graphical representation of zeta potential as a function of pH for P25 and the prepared TiO2 suspension doped with nitrogen in the presence of NaCl (10-3 M).

sl. 5. FT-IR spektri P25, TO-CNH3 i TO-PNH3.

to form lattice nitrogen. This is probably because the NH3 plasma consists of different nitrifying species, which more easily leads to the formation of lattice nitrogen [23]. In addition, traces of N species, located at 395.3 eV, were present in TO-PNH3 and TO-CNH3. Li et al. [24] prepared N-doped TiO2 in liquid NH3/ethanol under supercritical conditions and suggested that the N species with binding energy at 395.3 eV was attributed to surface adsorbed NH3 molecules. In this study, N species located at 395.3 eV existed only in TO-PNH3 and TO-CNH3, which were prepared using NH3 as the nitrogen source. This confirmed the result of Li et al. In addition, the peak intensity of TO-CNH3 at 395.3 eV was obviously higher than that of TO-PNH3, indicating that more NH3 molecules were adsorbed on the surface of TO-CNH3. Fig. Figure 4 shows the XP spectra of P25 and prepared N-doped TiO2 samples in the Ti2p and O1s regions. Compared to P25 spectra, clear shifts towards lower binding energies were observed for N-doped TiO2 samples in the Ti 2p region (458.4 eV) as well as the O1s region (529.7 eV). This is probably attributed to a change in the chemical environment after N-doping [24]. It is known that the binding energy of an element is influenced by its electron density. A decrease in binding energy implies an increase in electron density. Electrons in N atoms can be partially transferred from N to Ti and O due to the higher electronegativity of oxygen, leading to increased electron density on both Ti and O. The peaks around 530 and 532 eV in the O1s region are attributed to crystal lattice oxygen (Ti–O ) and the surface hydroxyl group (O–H) of TiO2. The ratio of these two peak areas (SO-H /STi-O) represents the abundance of surface hydroxyl groups. The calculated results showed that the SO-H/STi-O ratio of TO-CNH3 was 0.06, much lower than that of P25 (0.14). While the SO-H/STi-O ratios of TO-PN2 and TO-PNH3 were 0.13 and 0.11, which were slightly lower than P25. This showed that the content of surface hydroxyl groups decreased more drastically after the calcination process under NH3 current compared to the plasma treatment. It is known that these surface hydroxyl groups play an important role in photocatalysis. They react with photogenerated holes and produce active hydroxyl radicals, which are responsible for photodegradation [25]. FT-IR spectra of P25 and TO-PNH3 are shown in fig. 5. The absorption peak at 1620 cm−1 is attributed to the bending vibration of the hydroxyl group. The band at about 655 cm−1 belongs to the O-Ti-O structure of TiO2. There are three bands at 1402, 1224 and 1070 cm−1 observed in the spectra of TO–CNH3 and TO–PNH3, but not in P25. The band at 1402 cm−1 is attributed to surface adsorbed NH3 species on Brönsted acid sites (–OH)

[26]. It is known that NH3 can be adsorbed on Brönsted acid sites (–OH) located at 1400 cm−1 and Lewis acid sites (Ti4+) located at 1225 and 1190 cm−1 [26,27]. In Fig. 5, however, no NH3 adsorbed on the Lewis acid sites was observed. There are many previous literatures showing FT-IR results of NH3 adsorbed on TiO2 materials. Some of them reported that NH3 was adsorbed on Brönsted acid and Lewis acid sites [26,27]. Other results showed that adsorption was achieved only on Brönsted acid sites, which is consistent with the result in Fig. 5 [24,28]. Therefore, it is suggested that the preparation methods and conditions probably affect the adsorption state, leading to the adsorption site being different from different literatures. It is known that the surface of TiO2 is hydrophilic. In this study, TiO2 materials were treated under annealing and plasma conditions for only 15 minutes, resulting in most of the H2O adsorbed on Ti4+ still remaining on the TiO2 surface. It has been reported that when Ti4+ sites are saturated with hydroxyl groups, NH3 will mainly adsorb on Brönsted acid sites with the formation of N···HO bonds [29]. Therefore, only adsorption on Brönsted acid sites was achieved. In Fig. 5, the peaks at 1224 and 1070 cm−1 can be attributed to nitrogen atoms embedded in the TiO2 network, which is consistent with the XPS result [28]. These results confirmed the formation of N doping species in the TiO2 lattice. Figure 6 shows zeta-potential curves as a function of pH for P25 and prepared N-substituted TiO2 suspensions in the presence of NaCl (10-3 M). It is known that the point of zero charge (PZC) of TiO2 is around 3-6, which indicates that the surface of the TiO2 particles is positively charged. Compared to P25, clear shifts towards the lowest PZC value were observed for all nitrogen-doped TiO2, indicating that the positive charge on the TiO2 surface was reduced. The PZC value decreased in the following order: P25 > TO–PN2 > TO–PNH3 > TO–CNH3 . It is possible that the N-doping electron pair neutralizes some positive charges. In addition, NH3 was easily adsorbed on the catalyst surface during the nitration process due to the many acidic hydroxyl groups on the TiO2 surface. The presence of these adsorbed NH3 on the surface reduced the number of acidic hydroxyl groups, resulting in a lower PZC value of TO-PNH3 and TO-CNH3. In addition, plasma treatment caused a more drastic degradation of NH3, leading to less adsorbed NH3 on the surface of TO-PNH3 compared to TO-CNH3. Therefore, the PZC value of TO-PNH3 is higher than that of TO-CNH3. The adsorption of MB on TiO2-based catalysts was measured by the equilibrium adsorption capacity. The adsorption capacity of all TiO2 samples doped with nitrogen was lower than that of P25 (Figure 7). The specific BET surface area (SBET), pore volume, and mean pore size are listed in Table 1. Compared to P25, the SBET, pore volume, and mean pore size decreased for the as-prepared samples.

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100

TO-PNH3

TO-PNH3(HCl)

TO-PN2 4

HEALTHY/%

TO-PNH3 3

TO-CNH3

2 20

Fig. 7. MB adsorption capacity on P25 and N-doped TiO2 samples.

This probably caused the reduced equilibrium adsorption capacity shown in Fig. order: P25 > TO–PN2 > TO–PNH3 > TO–CNH3, which is completely consistent with the order of surface hydroxyl group content. This showed that the content of surface hydroxyl groups significantly affects the equilibrium adsorption capacity. The equilibrium adsorption capacities of TO-PNH3 and TO-CNH3 were shown to be lower than TO-PN2, which uses N2 as a nitrogen source. With the content of hydroxyl groups lower than TO-PN2, the large number of NH3 molecules adsorbed on the hydroxyl groups of TO-PNH3 and TO-CNH3 caused reduced surface sites for MB adsorption, which leads to a lower equilibrium adsorption capacity than TIL-PN2. The photocatalytic performance under visible light shown in Fig. 8 shows that the prepared TiO2 samples doped with N showed much higher activities than P25. Since no obvious changes in phase composition and particle size were observed between P25 and as-prepared N-doped TiO2 samples, the increased photocatalytic activity must be a consequence of nitrogen doping in TiO2, which led to a decrease in band gap and thus increased absorption in the visible region. In addition, the photocatalytic activity appears to increase in the order: TO–PN2 > TO–CNH3 > TO–PNH3, which is consistent with the order of lattice nitrogen content (Nfresh). This proved that lattice nitrogen was significantly 100

P25 TO-PN2

TO-CNH3 TO-PNH3

HEALTHY/%

TO-PNH3(H2O)

-6

Adsorption amount / 10 g/g

P25

0 0

t/h Sl. 8. Photocatalytic performance of P25 and the prepared N-doped TiO2 samples in the degradation of MB under visible light irradiation.

t/h Sl. 9. Photocatalytic performance of TO–PNH3, TO–PNH3 (H2O) and TO–PNH3 (HCl) in MB degradation under visible light irradiation.

affected visible light activity, which is consistent with Yamada's previous findings [10]. On the other hand, the stronger absorption of TO-PNH3 in the visible light region caused a more efficient use of visible light, leading to much higher activity than TO-PN2 and TO-CNH3. Nused is calculated and shown in Table 1. It is clear that Nused for TO-CNH3 and TO-PN2 is much lower than for Nfresh, while lattice nitrogen in TO-PNH3 is relatively stable. It was reported that lattice nitrogen was oxidized by photogenerated holes during the decomposition reaction, leading to a decrease in lattice nitrogen content [30]. Therefore, it is concluded that the oxidation of the TO-PNH3 lattice nitrogen is more difficult than in the other two samples. This difference in lattice nitrogen stability is likely due to the different preparation method between the three samples. Furthermore, Chen et al. [30] prepared N-doped TiO2 by heating TiO2 powder in NH3 flow and found that the presence of surface adsorbed NH3 reduced the number of surface sites available for reactants, resulting in low photocatalytic activity. In this study, compared to TO-PNH3, more NH3 was adsorbed on the surface of TO-CNH3, which led to a lower adsorption capacity and photocatalytic activity of TO-CNH3. To confirm the harmful effect of NH3, Chen et al. [30] washed the prepared N-doped TiO2 several times with pure water to remove adsorbed NH3. The photocatalytic activity of the obtained sample was improved after washing, but still significantly lower than that of the sample after calcination (NT400). This is probably because the NH3 was not completely removed by washing with clean water. In this study, TO-PNH3 was washed with HCl (0.1 M) to remove adsorbed NH3 and then cleaned with deionized water. The resulting sample was designated as TO-PNH3 (HCl). For comparison, TO-PNH3(H2O) was obtained by washing TO-PNH3 directly with deionized water. FT-IR results (not shown) showed that the adsorbed NH3 was completely removed after washing with HCl, while the remaining NH3 still remained on the TO-PNH3(H2O) surface. The photocatalytic performance (Fig. 9) shows that the activity increased in the order: TO–PNH3 < TO–PNH3 (H2O) < TO–PNH3 (HCl), which confirms the harmful effect of NH3 on the photocatalytic activity. When TO-PN2 and TO-CNH3 were used to replace TO-PNH3 following the same procedure as in the preparation of TO-PNH3(HCl), the product was recorded as TO-PN2(HCl) and TO-CNH3(HCl), resp. The photocatalytic efficiency of TO-PN2(HCl), TO-CNH3(HCl) and TO-PNH3(HCl) was tested in three cycles to examine the photocatalytic stability (Figure 10). It was shown that the activity of TO-PNH3 (HCl) decreased slightly in the 1st recycle and remained constant in the next two cycles. For TO–PN2 (HCl) i

S. Hu i sur. / Journal of Hazardous Materials 196 (2011) 248–254

253

120 110

TO-PNH3(HCl)

TO-CNH3(HCl)

N 1s

TO-PN2(HCl)

after deleting Ar

100 90

Intensity / a.u.

HEALTHY/%

80 70 60 50 40

TO-PNH3(HCl)

TO-CNH3(HCl)

20 10 0

TO-PN2(HCl) pretres

1. recycling

2. recycling

3. recycling

Fig. 10. Photocatalytic stability of the prepared N-doped TiO2 samples in the degradation of MB.

TO-CNH3 (HCl), the activities gradually decreased from 41.8% and 55.1% for the fresh catalyst to 30.9% and 49.6% for the 3rd recycled catalyst. This implies that the photocatalytic stability of TO-PNH3(HCl) is much better than TO-PN2(HCl) and TO-CNH3(HCl). This difference in photocatalytic stability is assumed to be attributed to the different lattice nitrogen stability between the three samples. Therefore, lattice nitrogen contents of fresh and recycled TO–PN2(HCl), TO–CNH3(HCl) and TO–PNH3(HCl) were calculated according to the relevant XPS data (Table 2). The lattice nitrogen content for TO-PN2(HCl) and TO-CNH3(HCl) gradually decreased from 1.28 at.% and 1.62 at.% to 0.46 at.% and 1.04 at.% after three cycles, which confirmed that the lattice- Nitrogen significantly affected the activity of visible light. However, the lattice nitrogen content of TO-PNH3 (HCl) decreased slightly from 1.94 to 1.78 to 1.78 in the 1st cycle and then remained constant in the next two cycles. This indicated that the lattice nitrogen atoms in TO-PNH3(HCl) remained relatively stable. As mentioned above, Nused for TO-CNH3 and TO-PN2 is much lower than for Nfresh, while lattice nitrogen for TO-PNH3 is stable (Table 1). This difference in lattice nitrogen stability is likely due to the different preparation method between the three samples. To elucidate why the photocatalytic stability is different between the samples, the XP spectra of fresh TO-PN2 (HCl), TO-CNH3 (HCl), and TO-PNH3 (HCl) in the N1s region after Ar+ ion etching were measured and shown in Fig. 11. Naturally, the NH3 species adsorbed on the surface located at 395.3 eV and the N-N (N-O) species located at 400 eV were removed after Ar + ion etching to remove the surface layer. For all three samples, only one peak around 396 eV, attributed to lattice nitrogen, was observed. Calculation according to the relevant XPS data showed that the lattice nitrogen contents for TO–PN2(HCl), TO–CNH3(HCl), and TO–PNH3(HCl) were 0.32, 0.74, and 1.58, respectively, at .% Compared with fresh catalyst data in Table 2, more than 50% and 75% of lattice nitrogen in TO-CNH3 (HCl) and TO-PN2 (HCl) were eliminated after Ar + ion etching. This showed that a large number of N atoms were doped only in the surface layer of TO-CNH3(HCl) and TO-PN2(HCl), which were easily oxidized by photogenerated holes. ( HCl), TO–CNH3 (HCl), and TO–PN2 (HCl) determined from XPS data. A sample

Fresh catalyst (in %)

1. recycling (in %)

2. recycling (in %)

3. recycling (in %)

TO–PNH3 (HCl) TO–CNH3 (HCl) TO–PN2 (HCl)

1,94 1,62 1,28

1,78 1,31 1,02

1,78 1,22 0,75

1,75 1,04 0,46

390

395

400

405

Binding energy / a.u. Fig. 11. XP spectra of TO-PN2 (HCl), TO-CNH3 (HCl) and TO-PNH3 (HCl) in the N1s region after Ar + ion etching.

Table 3. Comparison of nitrogen stability in the lattice of N-doped TiO2 samples before and after washing with HCl solution. Sample

Fresh catalyst (in %)

1. recycling (in %)

Retention rate

TO–PNH3 (HCl) TO–CNH3 (HCl) TO–PN2 (HCl) TO–PNH3 TO–CNH3 TO–PN2

1,94 1,62 1,28 1,95 1,64 1,32

1,78 1,31 1,02 1,91 1,17 0,76

0,92 0,81 0,80 0,98 0,71 0,58

The retention rate is equal to the ratio of the lattice nitrogen content of the first recycled catalyst to the fresh catalyst.

during the decomposition reaction, which leads to a decrease in the nitrogen content in the network. Therefore, TO-PN2 and TO-CNH3 showed poor photocatalytic stability. In contrast, compared to the data for fresh TO-PNH3(HCl) in Table 2, less than 20% of the lattice nitrogen of TO-PNH3(HCl) was removed after Ar + ion etching. This is probably because the excited hydrogen species produced by the NH3 plasma caused impregnation of nitrogen atoms into the crystal lattice of the deeper layer, causing difficulties in the oxidation of photogenerated holes. Therefore, the photocatalytic stability of TO-PNH3(HCl) was much higher than that of TO-PN2 and TO-CNH3. The lattice nitrogen retention rate, which represents the lattice nitrogen stability, was calculated and shown in Table 3. Compared to the sample before rinsing with HCl, more than 10% and 20% increase in the retention rate was observed in TO-CNH3 (HCl) and TO–PN2 (HCl), while in TO–PNH3 (HCl) only a small decrease in the retention rate occurs. This showed that the lattice nitrogen stability of N-doped TiO2 samples was improved after rinsing with HCl solution. This is likely because the surface N species adsorbed on the Brönsted acid sites (–OH) are removed by the HCl solution, making more surface hydroxy groups available to capture photogenerated holes, limiting nitrogen oxidation in the photohole lattice. 4. Conclusion NH3 plasma, N2 plasma and NH3 liquid annealing were used to prepare N-doped TiO2, respectively, to investigate the effect of preparation method, nitrogen source and post-treatment on photocatalytic activity and stability. The photocatalytic activity increased in the order: TO-PN2 < TO-CNH3 < TO-PNH3, indicating that NH3 plasma is more effective among the three

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methods. Lattice nitrogen significantly affected visible light activity. NH3 adsorbed on the catalyst surface led to a low adsorption capacity of reactive MB, which resulted in reduced photocatalytic activity. After removal of NH3 by washing with HCl, the obtained catalysts showed significantly higher activity under visible light, confirming the harmful effect of NH3. The photocatalytic stability of N-doped TiO2 prepared by NH3 plasma was much higher than that of samples prepared by other nitration procedures. This suggested that the excited hydrogen species produced by the NH3 plasma caused the absorption of N atoms into a deeper layer of the crystal lattice, which hindered oxidation by photogenerated holes from other catalysts, leading to stable nitrogen in the lattice. Moreover, the lattice nitrogen stability of N-doped TiO2 samples was improved after rinsing with HCl solution. This is likely because the surface N species adsorbed on the Brönsted acid sites (–OH) were removed by the HCl solution, resulting in more surface hydroxyl groups to trap photogenerated holes, thus limiting lattice nitrogen oxidation. This stable lattice nitrogen during the decomposition reaction caused high photocatalytic stability. Acknowledgments This work was supported by National Natural Science Foundation of China (No. 41071317, 30972418), National Key Technology Research and Development Program of China (No. 2007BAC16B07), Natural Science Foundation of Liaoning Province (No. 20092080). The authors would like to thank prof. Anjie Wang, Dalian University of Technology, for their contributions to the manuscript. Literature [1] A. Fujishima, T.N. Rao, D.A. Press, Titanium dioxide photocatalysis, J. Photochem. Photobiol. C 1 (2000) 1-21. [2] M.R. Hoffmann, S.T. Martin, W. Choi, D.W. Bahnemann, Applications of semiconductor photocatalysis in the environment, Chem. Rev. 95 (1995) 69-96. [3] R. Asahi, T. Morikawa, T. Ohwaki, A. Aoki, Y. Taga, Visible light photocatalysis in nitrogen-doped titanium oxides, Science 293 (2001) 269-271. [4] T. Lindgren, J.M. Mwabora, E. Avendano, J. Jonsson, A. Hoel, C.G. Granqvist, S.E. Lindquist, Photoelectrochemical and optical properties of nitrogen-doped titanium dioxide films prepared by DC reactive magnetron sputtering, J. Phys. Chem. B 107 (2003) 5709-5716. [5] M. Qiao, S.S. Wu, Q. Chen, J. Shen, Novel triethanolamine-assisted sol-gel synthesis of N-doped TiO2 hollow spheres, Mater. Easy. 12 (2010) 1398-1400. [6] H. Shen, L. Mi, P. Xu, W.D. Shen, R.N. Wang, Visible light photocatalysis of nitrogen-doped TiO2 nanoparticle films prepared by low-energy ion implantation, Appl. Surf. Sci. 17 (2007) 7024-7028. [7] L. Zhao, Q. Jiang, J.S. Lian, Visible light photocatalytic activity of nitrogen-doped TiO2 thin film prepared by pulsed laser deposition, Appl. Surf. Sci. 15 (2008) 4620-4625. [8] S.Z. Hu, A.J. Wang, X. Li, H. Löwe, Hydrothermal synthesis of well-dispersed ultran-N-doped TiO2 nanoparticles with enhanced photocatalytic activity under visible light, J. Phys. Chem. Solids 71 (2010) 156–162.

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Journal of Hazardous Materials 196 (2011) 255-262

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Enhanced photocatalytic activity of Bi2 WO6 with oxygen vacancies doped with zirconia Zhijie Zhang, Wenzhong Wang ∗ , Erping Gao, Meng Shang, Jiehui Xu State Key Laboratory of High-Performance Ceramics and Superfine Microstructures, Shanghai Academy of Sciences, Shanghai 1295 Road, Shanghai 200 , LD China

article

i n f o

Article history: Received July 3, 2011 Received in revised form September 5, 2011 Accepted September 6, 2011 Available online September 10, 2011 Keywords: Zr4+ -doped Bi2 WO6 Oxygen vacancy Photocatalysis RhB Phenol

a b s t r a c t In order to overcome the drawback of low photocatalytic efficiency caused by electron-hole recombination, zirconia-doped Bi2 WO6 photocatalysts with an oxygen vacancy were synthesized. Oxygen vacancies as centers of positive charge can easily capture an electron, thus inhibiting the recombination of charge carriers and extending the electron lifetime. In addition, the creation of oxygen vacancies favors the adsorption of O2 on the semiconductor surface, which facilitates the reduction of O2 by trapped electrons to form peroxide radicals, which play a key role in the oxidation of organic substances. Visible light-induced photodegradation of rhodamine B (RhB) and phenol was performed to evaluate the photoactivity of the products. The results showed that the oxygen-poor Bi2WO6 exhibited much better photoactivity than the oxygen-deficient Bi2WO6 photocatalyst. This work provided a new idea for the rational design and development of high-performance photocatalysts. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Semiconductor-based photocatalysis has attracted much attention due to its potential applications in renewable energy and environmental fields, such as dye-sensitized solar cells, hydrogen production from water splitting, and photocatalytic water/air purification [1-5]. Since these applications rely on the photogeneration of charge carriers such as electrons and holes, the success of the applications depends on the efficiency of electron or hole transfer, which is closely related to the recombination rates of the photogenerated charge carriers. Due to the much higher recombination rate (nanoseconds) than the interfacial transfer rate (microseconds to milliseconds), many charge carriers recombine and dissipate the input energy as heat, greatly limiting the overall quantum efficiency of photocatalysis [6]. Therefore, to improve the photocatalytic activity of semiconductors, it is important to control the recombination dynamics of photogenerated charge carriers. If a suitable scavenger or surface defect state is available to trap an electron or hole, recombination is inhibited and a subsequent redox reaction can occur. It has been reported that oxygen vacancies can act as electron capture centers and therefore play an important role in delaying the recombination of charge carriers, which can lead to increased photocatalytic activity of photocatalysts [7-9]. Furthermore, existing oxygen vacancies

they can significantly act as specific reaction sites for reactant molecules in heterogeneous reactions [10]. Therefore, the introduction of oxygen vacancies into photocatalysts could be a feasible approach for the development of highly active photocatalysts. As one of the simplest Aurivillius oxides with a layered structure, Bi2 WO6 has recently attracted considerable attention due to its good photocatalytic properties in water splitting and organic pollutant degradation under visible light irradiation [11-15]. So far, much has been done to facilitate electron-hole separation and improve the photocatalytic activity of Bi2WO6, including surface modification [16,17], anionic doping [18], and coupling with other semiconductors [19,20]. As far as we know, the effect of oxygen vacancies on the photocatalytic activity of Bi2WO6 has rarely been reported. Here, for the first time, we introduce oxygen vacancies in Bi2WO6 by doping with zirconium, and the correlation between oxygen vacancies and the photocatalytic activity of Bi2WO6 is also investigated. Bi2WO6 does not contain oxygen vacancies, and it has been reported that substitution of W with corresponding cations with lower valence states could lead to exogenous oxygen deficiency through charge compensation [21,22]. Error calculations show that the low solvation energy (0.05 eV) is favorable for the substitution of ZrIV at the WVI position by creating external oxygen vacancies [22], which can be described by defect reactions written as: WO

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3 ZrO2 −→Zr

∗ Corresponding author. Phone: +86 21 5241 5295; fax: +86 21 5241 3122. E-mail address:[email protected](W. Wang). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.017

+ VO •• +

In this research, we succeeded in preparing oxygen-deficient Bi2 WO6 phases by substituting WVI with ZrIV. Photo activity

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Z. Zhang i sur. / Journal of Hazardous Materials 196 (2011) 255-262

Fig. 1. (A) XRD patterns of the synthesized products. (B) positions of the diffraction peaks for the plane (1 3 1) in the range 2 = 27.5–29◦ .

Evaluation by photocatalytic decomposition of RhB and phenol under visible light showed that the photocatalytic activity depends on the oxygen vacancy concentration, and Zr4+-doped Bi2WO6 shows much better photocatalytic performance than the undoped Bi2WO6 sample. Furthermore, the role of the oxygen vacancy in promoting the separation of charge carriers and improving the photocatalytic activities was elucidated in detail.

2. Experimental 2.1. Preparation of Zr4+-doped Bi2WO6 photocatalysts Zr4+-doped Bi2WO6 photocatalysts were prepared by the hydrothermal method. In a typical procedure, 2 mmol of Bi(NO3)3·5H2O and 1 mmol of Na2WO4·2H2O were dissolved in 2 mL of 2 M nitric acid and 30 mL of deionized water, respectively. Then these two solutions were mixed together and stirred for 30 minutes. Then an aqueous solution containing the desired amounts of ZrOCl2·8H2O for Bi2WO6 doped with Zr4+ was added. The molar ratios between Zr and Bi2WO6 were set to 0, 2.0%, 3.0%, and 4.0%, respectively, and the corresponding products were named Zr-0, Zr-0.02, Zr0.03, and Zr-0, 04. The pH of the final suspension was adjusted to approx. 7 and the mixture was stirred for several hours at room temperature. The suspensions were then added to a 50 ml Teflon-lined autoclave to 80% of the total volume. The autoclave was sealed in a stainless steel container and heated to 160 ◦ C for 24 hours. The autoclave was then naturally cooled to room temperature. The products were collected by filtration, washed several times with distilled water and then dried at 60 ◦ C in air for 12 hours.

2.2. Characterization The phase and composition of the prepared samples were measured by X-ray diffraction (XRD) using an X-ray diffractometer with Cu K ␣ radiation under 40 kV and 100 mA and with 2 in the range of 20◦ to 60◦ (Rigaku, Japan). The morphologies and microstructures of the prepared samples were examined by transmission electron microscopy (TEM, JEOL JEM-2100F). UV-vis diffuse reflectance spectra (DRS) of the samples were recorded with a UV-vis spectrophotometer (Hitachi U-3010) using BaSO4 as a reference. The chemical compositions of the manufactured products were analyzed by X-ray photoelectron spectroscopy (XPS) analysis (Thermo Scientic Escalab 250). All binding energies are referenced to the C1s peak (284.8 eV) originating from the random carbon. The photoluminescence (PL) spectra of the samples were recorded

with Perkin Elmer LS55. Analysis of total organic carbon (TOC) was performed with a liquid TOC II elemental analyzer. 2.3. Photocurrent measurement Photocurrent measurements were performed using a CHI 660C electrochemical workstation. 25 mg of photocatalyst was suspended in deionized water (50 mL) containing acetate (0.1 M) and Fe 3+ (0.1 mM) as electron donor and acceptor. A Pt plate (both sides exposed to the solution), saturated calomel electrode (SCE) and Pt gauze were immersed in the reactor as working (collector), reference and counter electrodes, respectively. Photocurrents were measured by applying a potential (+1 V vs. SCE) to the Pt electrode using a potentiostat (EG&G). 2.4. Measurement of photocatalytic activities The photocatalytic activities of Bi2 WO6 photocatalysts doped with Zr4+ were measured by monitoring the photodegradation of rhodamine B (RhB) and phenol in aqueous solution. 100 mg of photocatalyst was dispersed in 100 mL of RhB solution (10-5 mol/L) or phenol (20 mg/L). Before illumination, the suspensions were magnetically stirred in the dark for 1 h to ensure adsorption/desorption equilibrium of RhB or phenol with the photocatalyst powders and then exposed to visible light from a 500 W Xe lamp with a 420 nm cutoff filter. After a certain time of irradiation, a 3 ml sample of the suspension was taken and centrifuged to remove the photocatalysts. The supernatant was then removed to measure the spectral change in absorbance of RhB or phenol using a UV-vis spectrophotometer (Hitachi U-3010) to monitor the rate of photodegradation. The change in the concentration of rhodamine B and phenol was determined by monitoring the optical intensity of the absorption spectra at 553 nm and 270 nm, respectively. 3. Results and discussion 3.1. Crystal structure and product morphology XRD diffractograms of pure Bi2 WO6 and Zr4+ doped Bi2 WO6 samples are shown in Figure 1(A). All diffraction peaks correspond to typical data for the structure of Russellite Bi2WO6 (JCPDS 39-0256) and no characteristic peaks of any impurities are detected in the patterns, indicating that zirconium doping does not lead to the growth of new phases. However, a careful comparison of the (1 3 1) diffraction peaks in the region 2 = 27.5–29◦ (Fig. 1(B)) shows that the peak position of Bi2 WO6 shifts slightly to a lower value of 2 with increasing

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sl. 2. TEM micrograph of (A) Zr-0 and (B) Zr-0.03.

zirconium content. The same results are shown in other diffraction peaks. According to Bragg's law, d(h k l) = /(2 sin ), where d(h k l) is the distance between the crystal planes (h k l), the wavelength of the X-ray and the diffraction angle of the crystal plane. (h k l ) [23], must be reduced by a value of 2 due to an increase in lattice parameters (value d(1 3 1)). Since the ionic radius of Zr4+ (0.080 nm) is smaller than that of Bi3+ (0.108 nm) but larger than that of W6+ (0.062 nm), the observed shift of the diffraction peak to lower angles should be due to the expected larger lattice parameter for replacing W6+ with Zr4+. To obtain detailed information on the microstructure and morphology of the synthesized samples, TEM observations

is carried out. Figure 2(A) and (B) show representative TEM images of pure Bi2 WO6 sample and Bi2 WO6 sample with 3.0 mol% zirconium content. Both samples show plate-like morphology, indicating that zirconium doping has no obvious effect on the morphology of Bi2WO6. 3.2. X-ray photoelectron spectroscopy (XPS) analysis The surface composition and elemental oxidation states of the sample prepared with 3.0 mol% zirconium content were investigated by XPS analysis, and the corresponding experimental results are shown in the figure. 3. Total XPS spectra shown in fig. 3(A)

sl. 3. XPS spectra of Zr-0.03. (A) Total XPS spectrum of the sample. (B) Bi4f spectrum. (C) W 4f spectrum i (D) Zr 3d spectrum.

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Fig. 4. High-resolution O1s XPS spectra of (A) Zr-0 and (B) Zr-0.03.

indicates that all the peaks in the curve are attributed to the elements Bi, W, O, Zr and C, and no peaks of other elements are observed. The presence of C comes mainly from the carbon films used for XPS measurement. Parts B-D in Fig. 3 show the high-resolution spectrum for Bi, W and Zr species. According to Fig. 3(B), the binding energies of Bi 4f7/2 and Bi 4f5/2 are 159.2 eV and 164.4 eV, respectively, which corresponds to the characteristic peak of Bi3+. The W 4f orbital is clearly split into W 4f5/2 and W 4f7/2 contributions, centered at 37.3 eV and 35.2 eV, respectively (Fig. 3(C)), which are very close to previously reported values [ 19 ] , suggesting that the tungsten in the Zr4+ doped Bi2 WO6 sample exists as W6+. The Zr 3d spectra shown in Fig. 3(D) consists of the main peaks Zr 3d3/2 and Zr 3d5/2 with a peak spacing of 2.4 eV, which is consistent with the literature data for Zr4+ [24]. In addition, we investigated the presence of oxygen vacancies from XPS spectra. The high-resolution O 1s XPS spectra of Bi2 WO6 and Zr4+ -doped Bi2 WO6 samples without doping are shown in Fig. 4(A) and (B). Both profiles are asymmetric and can be fitted by two Gaussian features, usually assigned as a low binding energy component (LBEC) and a high binding energy component (HBEC), indicating two different types of O species in the sample. LBEC and HBEC can be attributed to lattice oxygen and chemisorbed oxygen caused by surface chemisorbed species such as hydroxyl and HO [ 25 ]. It was previously reported that the HBEC component increases with increasing oxygen vacancy [ 26 ], which may lead to an asymmetry of the main peak. High-resolution O 1s XPS spectra show that the HBEC peak area is clearly larger in the zirconium-swollen sample compared to the uncoated sample. In addition, the calculated adsorbed oxygen to lattice oxygen ratios are 1.14 and 0.43 for the zirconia-doped sample and the undoped sample, respectively, strongly indicating the presence of oxygen deficiency in the zirconia-doped sample.

associated with oxygen vacancies just below the conduction band minimum [27]. Oxygen vacancies are positive charge centers that easily bind electrons. Excitation of electrons from such localized states in the conduction band can lead to better absorption of visible light. Therefore, with a higher zirconium concentration, more oxygen vacancies are created and the optical absorption properties of the samples become stronger. 3.4. Photoluminescence spectra and photoelectrochemical measurements Since photoluminescence (PL) emission mainly originates from the recombination of free carriers, PL spectra are useful for determining the migration, transport and recombination processes of photogenerated electron-hole pairs in a semiconductor. The weaker PL intensity implies a low rate of electron-hole recombination under light irradiation [28]. Figure 6(A) shows the PL spectra of undoped Bi2WO6 and Bi2WO6 with incorporated Zr4+ (Zr-0.03) when the excitation wavelength was 300 nm. There was a significant decrease in the intensity of the PL spectrum of Zr4+-doped Bi2WO6, confirming that zirconium doping can effectively inhibit the recombination of photogenerated charge carriers. Photocurrent indirectly reflects the semiconductor's ability to generate and transport photogenerated charge carriers, which correlates with photocatalytic activity [29]. In order to investigate the photoinduced charge separation performance of the uncoated and zirconium oxide samples, photocurrent measurements were performed under visible light irradiation. As shown in Figure 6(B), the Zr0.03 sample produces a higher photocurrent than the non-regenerated Bi2 WO6,

3.3. Optical properties of the product. UV diffuse reflectance spectra (DRS) of Zr4+ doped Bi2 WO6 samples compared to pure Bi2 WO6 are shown in the figure. 5. A sample of pure Bi2 WO6 showed photoabsorption from the UV light range to visible light with a wavelength less than 450 nm. It was noteworthy that the absorption onset of Zr4+-doped Bi2 WO6 samples was clearly red-shifted. With the increase in the amount of zirconium doping, the intensity of visible light absorption of the samples became stronger, which can be attributed to the excitation of trapped electrons in localized states

Fig. 5. UV-vis diffuse reflection spectra of the produced samples.

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Fig. 6. (A) Photoluminescence spectrum at room temperature (PL) of Zr-0.03 and Zr-0 (Ex = 300 nm). (B) photocurrent generated with visible light irradiation time over Zr-0.03 and Zr-0 suspended with acetate and Fe3+.

showing that zirconium doping can effectively enhance charge carrier transport and reduce electron-hole recombination. 3.5. Photocatalytic performance of Zr4+-doped Bi2 WO6 RhB samples, a hazardous compound as well as a common pollutant, was selected as a representative pollutant to evaluate the photocatalytic performance of the photocatalyst. The variation of RhB concentration with respect to reaction time in the presence of Zr4+ doped Bi2WO6 samples compared to pure Bi2WO6 is illustrated in the figure. 7(A). The results show that the photoactivity of the samples was highly dependent on the zirconium doping concentration. As the zirconium concentration increased from 0 mol% to 3.0 mol%, the photoactivity of the samples for RhB photodegradation improved. When the zirconium concentration increased to 4.0 mol%, the photoactivity decreased compared to that of 3.0 mol%, but still higher than that of undeveloped Bi2WO6. The highest photoactivity was observed for Zr-0.03, which can completely degrade RhB in 20 minutes, while only 65.4% of RhB was degraded in the presence of pure Bi2 WO6 in the same time period. Furthermore, a comparison of the apparent rate constant k in Fig. 7(B) shows that Zr-0.03 had the highest value of k in the photodegradation of RhB, while that of Zr-0.04 decreased compared to that of Zr-0.03. The reason can be interpreted as follows: an adequate amount of oxygen vacancies can trap electrons, causing holes to freely diffuse to the surface of the semiconductor, where oxidation of organic substances can occur.

Therefore, the appropriate content of oxygen vacancies will improve the photocatalytic process by efficiently separating electron-hole pairs. However, if it exceeded the optimal value, the oxygen vacancies would act as recombination centers for photoinduced electrons and holes, which is unfavorable for photocatalytic performance [27,30]. When the concentration of Zr doping was 3.0 mol%, the corresponding content of oxygen vacancies was generated and this photocatalyst showed the highest photocatalytic activity. When the Zr doping concentration was further increased, the excessive oxygen vacancies created led to poor photocatalytic performance. Therefore, an appropriate amount of zirconium doping can significantly improve the photocatalytic activity of Bi2WO6. The photocatalytic activities of the above-mentioned photocatalysts can be further controlled by the decomposition of another organic compound, such as phenol, which has no light absorption properties in the visible light region and no photosensitization, as shown in FIG. 8. It is clear when visible - Under light irradiation, phenol is more efficiently degraded by Zr-0.03 than by pure Bi2 WO6 (Fig. 8(A)). Approx. Degradation efficiencies of 62.5% and 14.2% were obtained within 120 minutes of Zr-0.03 and pure Bi2WO6, respectively. Due to the pseudo-first-order kinetics of phenol photodegradation in Bi2WO6, the apparent rate constant k was calculated to be 0.0012 min−1 and 0.0083 min−1 for pure Bi2WO6 and Zr-0.03, respectively (Figure (B) ). In other words, the photocatalytic activity of Zr-0.03 is about 7 times higher than pure Bi2WO6. To further investigate the photodegradation of phenol, total organic carbon (TOC), which is widely used to est

Fig. 7. (A) Photocatalytic degradation of RhB under visible light (> 420 nm) as a function of irradiation time of the prepared samples. (B) comparison of the rate constant k.

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Fig. 8. (A) Photocatalytic degradation of phenol under visible light irradiation with Zr-0.03 and Zr-0. (B) comparison of rate constant k; (C) TOC removal efficiency during the photocatalytic degradation of phenol in the presence of Zr-0.03 and Zr-0, respectively. (D) cycle in the photocatalytic degradation of phenol under visible light irradiation.

the degree of mineralization of the organic species was measured in the photodegradation process from the prepared samples under visible light, as shown in Fig. 8(C). The results confirm that phenol in the prepared samples is stably mineralized. Furthermore, the TOC removal efficiency in the presence of Zr-0.03 reaches a value of 29.2% after 120 minutes of irradiation, while for Bi2 WO6 it is only 5.3%. Based on the above results, it can be concluded that Zr-0.03 is a much better photocatalyst than pure Bi2WO6. To test the stability of the Zr4+-doped Bi2 WO6 photocatalyst, circulation pathways in the photocatalytic degradation of phenol were performed under visible light. As shown in Fig. In Fig. 8(D), after five cycles of phenol photodegradation, the catalyst showed no significant loss of activity, confirming that Zr4+-doped Bi2 WO6 was not photocorroded during the photocatalytic oxidation of pollutants. it is especially important for its implementation.

groups or H2O to form surface-bound hydroxyl radicals (• OH), and conduction band electrons can interact with adsorbed O2 to form peroxide radicals (O2 •−), which are also strong oxidizing species and play a key role in the oxidative degradation of organics [32, 33]. However, for the Bi2 WO6 system, holes could not react with OH− /H2O to form • OH due to the more negative redox potential of BiV /BiIII (+1.59 V) than • OH/OH− (+1.99 V) [34 ]. In order to determine the active species in the degradation process, holes and hydroxyl radical scavengers were added to the degradation system. Fig. 9 showed that the addition of isopropanol (IPA) as scavenger of hydroxyl radicals [34] caused little change in the photocatalytic degradation of phenol, indicating that • OH is not the main oxidation species in this process. However, when the hole cleaner EDTA [35] was introduced, the rate of phenol degradation

3.6. Mechanism of Enhanced Photoactivities Photocatalysis generally involves four processes [31]: (i) light-induced generation of conduction band electrons and valence band holes; (ii) transfer of photogenerated charge carriers to the surface of the photocatalyst. (iii) subsequent reduction/oxidation of adsorbed reactants directly by electrons/holes or indirectly by reactive oxygen species and (iv) recombination of photogenerated electron-hole pairs. The efficiency of photocatalysis is determined by the competition between the charge separation process and the charge recombination process. The desired photocatalysts are expected to promote charge transfer processes while suppressing the recombination process. For photocatalysts in aqueous solution, photoinduced valence band holes can usually react with hydroxyl atoms.

Fig. 9. Photocatalytic decomposition of phenol with the addition of cavity scavengers and hydroxyl radicals and in solutions saturated with N2 under visible light irradiation (> 420 nm).

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more peroxide anions on the photocatalyst surface. The generation of superoxide anions was a photoinduced electron capture process, which facilitated charge separation and led to a lower electron-hole recombination rate. Our work suggests that the idea of introducing oxygen vacancies could be an acceptable strategy for the development of efficient visible light photocatalysts for environmental remediation. Thank you

Fig. 10. Proposed photocatalytic mechanism of Bi2WO6 for oxygen deficiency. OP: organic pollutant? DP: degradation product.

they were in great depression. Therefore, holes play an important role in Bi2WO6 photocatalysis. The superoxide radical is another important intermediate for the oxidative degradation of organic substances [36,37]. In order to investigate the role of the superoxide radical in photocatalysis, the photocatalytic decomposition of phenol was carried out under conditions saturated with nitrogen. The result shown in Fig. 9 showed that under anoxic conditions the rate of phenol photodegradation was strongly suppressed, suggesting that the superoxide radical is an important oxidizing species in the photocatalytic process. Oxygen molecules as electron scavengers play a key role in photocatalysis by reacting with electrons to generate superoxide radicals. However, when the rate of O2 reduction by electrons is not fast enough to match the hole reaction rate, excess electrons will accumulate on the photocatalyst particles and the rate of electron-hole recombination will increase accordingly. In this case, electron transfer to O2 can be the limiting step in photocatalysis [38,39]. However, this can be overcome by introducing oxygen vacancies into the photocatalyst. In order to facilitate the reaction between oxygen and electrons, a strong adsorption of oxygen on the surface of the photocatalyst and a longer electron lifetime are required. Adsorption of O2 molecules has been reported to be mainly mediated by oxygen vacancies and oxygen physisorbers on defect-free oxide surfaces, but strongly interacts with oxygen vacancies [40,41]. Oxygen adsorption can therefore be enhanced by the formation of oxygen voids. On the other hand, if the electron moves freely in a semiconductor particle, it is difficult to avoid the fate of recombination. However, this can be avoided if the electrons are temporarily but effectively trapped in the particles. Positively charged VO •• defects can act as electron acceptors and can temporarily trap photogenerated electrons to reduce surface recombination of electrons and holes [8,9]. Therefore, the generation of oxygen vacancies can not only favor oxygen adsorption, but also delay the recombination of charge carriers, which facilitates the reduction rate of O2 to generate more superoxide anions (O2 •− ) on the photocatalyst surface (Fig. 10) and thus leads to increased photocatalytic activity. 4. Conclusions Bi2 WO6 photocatalysts with oxygen deficiency were synthesized by zirconium doping, and the relationship between oxygen vacancies and the photocatalytic activity of Bi2 WO6 was investigated. Visible light-induced photodegradation of RhB and phenol showed that zirconium doping can significantly improve the photocatalytic performance of Bi2WO6. The higher photocatalytic activity was attributed to the generation of oxygen vacancies, which promote the adsorption of O2 and the rate of O2 reduction, leading to

We gratefully acknowledge the financial support of the National Natural Science Foundation of China (50972155, 50902144, 50732004), the National Key Research Program of China (2010CB933503), and the Science Foundation for Young Scientists of the State Key Laboratory of High-Performance Ceramics Micro0L04 ). Literature [1] A.L. Linsebigler, G.Q. Lu, J.T. Yates, Photocatalysis on TiO2 surfaces: Basic mechanisms and selected results, Chem. Rev. 95 (1995) 735-758. [2] A. Hagfeldt, M. Gratzel, Light-induced redox reactions in nanocrystalline systems, Chem. Rev. 95 (1995) 49-68. [3] X.B. Chen, S.S. Mao, Titanium Dioxide Nanomaterials: Synthesis, Properties, Modifications and Applications, Chem. Rev. 107 (2007) 2891-2959. [4] W. Morales, M. Cason, O. Aina, N.R. Tacconi, K. Rajeshwar, Combustion synthesis and characterization of nanocrystalline WO3, J. Am. Chem. Soc. 130 (2008) 6318-6319. [5] W.J. Youngblood, S.H.A. Lee, K. Maeda, T.E. Mallouk, Visible light splitting of water using dye-sensitized oxide semiconductors, Acc. Chem. Res. 42 (2009) 1966-1973. [6] M.R. Hoffmann, S.T. Martin, W.Y. Choi, D.W. Bahnemann, Applications of semiconductor photocatalysis in the environment, Chem. Rev. 95 (1995) 69-96. [7] T.J. Kuo, C.N. Lin, C.L. Kuo, M.H. Huang, Growth of ultralong ZnO nanowires on silicon substrates by vapor transfer and their use as reusable photocatalysts, Chem. Mater. 19 (2007) 5143-5147. [8] J.C. Wang, P. Liu, X.Z. Fu, Z.H. Li, W. Han, X.X. Wang, Relationship between oxygen defects and photocatalytic properties of ZnO nanocrystals in Nafion films, Langmuir 25 (2009) 1218-1223. [9] Y.X. Zheng, C.Q. Chen, Y.Y. Zhan, X.Y. Lin, Q. Zheng, K.M. Wei, J.F. Zhu, Y.J. Zhu, Luminescence and photocatalytic activity of ZnO nanocrystals: structure-property correlation, Inorg. Chem. 46 (2007) 6675-6682. [10] X.Q. Gong, A. Selloni, M. Batzill, U. Diebold, Steps on anatase TiO2 (1 0 1), Nat. Mater. 5 (2006) 665-670. [11] A. Kudo, S. Hijii, Evolution of H2 or O2 from aqueous solutions on layered oxide photocatalysts composed of Bi3+ with 6s2 configuration and d0 transition metal ions, Chem. Easy. 28 (1999) 1103-1104. [12] C. Zhang, Y.F. Zhu, Synthesis of Square Bi2 WO6 Nanoplates as Photocatalysts with High Visible Light Activity, Chem. Mater. 17 (2005) 3537-3545. [13] H.B. Fu, L.W. Zhang, W.Q. Yao, Y.F. Zhu, Nanosized Bi2 WO6 catalysts prepared by hydrothermal synthesis and their photocatalytic properties, Appl. A cat. B: Environment. 66 (2006) 100-110. [14] F. Amano, A. Yamakata, K. Nogami, M. Osawa, B. Ohtani, Visible light-responsive pristine metal oxide photocatalyst: activity enhancement by crystallization under hydrothermal treatment, J. Am. Chem. Soc. 130 (2008) 17650-17651. [15] G.S. Lee, D.Q. Zhang, J.C. Yu, M.K.H. Leung, An efficient visible-light bismuth-tungstate photocatalyst for the decomposition of nitric oxide, Environ. Sci. Technol. 44 (2010) 4276-4281. [16] J. Ren, W.Z. Wang, S.M. Sun, L. Zhang, J. Chang, Enhanced photocatalytic activity of Bi2 WO6 loaded with Ag nanoparticles under visible light irradiation, Appl. A cat. B: Environment. 92 (2009) 50–55. [17] Y.Y. Lee, J.P. Liu, X.T. Huang, J.G. Yu, Carbon-modified Bi2 WO6 nanostructures with enhanced visible light photocatalytic activity, Dalton Trans. 39 (2010) 3420-3425. [18] M. Shang, W.Z. Wang, L. Zhang, H.L. Xu, Bi2 WO6 with significantly enhanced photocatalytic activities by nitrogen doping, Mater. Chem. Phys. 120 (2010) 155-159. [19] Q. Xiao, J. Zhang, C. Xiao, X.K. Tanning, photocatalytic degradation of methylene blue over Co3 O4/Bi2 WO6 composite under visible light irradiation, Catal. Commune. 9 (2008) 1247-1253. [20] L.S. Zhang, K.H. Wong, Z.G. Chen, J.C. Yu, J.C. Zhao, C. Hu, C.Y. Chan, P.K. Wong, AgBr–Ag–Bi2 WO6 nanojunction system: a novel and efficient photocatalyst with dual visible light active components, Appl. A cat. A 363 (2009) 221-229. [21] C.K. Lee, L.T. Sim, A.M. Kaput, A.R. West, On the possible doping of Bi2WO6 with Cu, J. Mater. Chem. 11 (2001) 1096-1099. [22] M.S. Islam, S. Lazure, R.N. Vannier, G. Nowogrocki, G. Mairesse, Structural and computational studies of oxygen ion conductors based on Bi2WO6, J. Mater. Chem. 8 (1998) 655-660. [23] G.L. Huang, Y.F. Zhu, Enhanced photocatalytic activity of ZnWO4 catalysts by fluorine doping, J. Phys. Chem. C 111 (2007) 11952-11958.

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Monitoring of PCBs in facilities associated with products and wastes containing PCBs in South Korea Guang-Zhu Jin a,1, Ming-Liang Fang a, Jung-Ho Kang a, Hyokeun Park a, Sang-Hyup Lee b, Yoon - Seok Chang a , b, * a b

School of Environmental Science and Engineering, Pohang University of Science and Technology (POSTECH), San 31, Hyoja-dong, Nam-gu, Pohang 790-784, Republic of Korea Water Research Center, Korea Institute of Science and Technology (KIST), Hwarangno, 14-Gil 5, Seongbuk-gu, Seoul 136-791, Republikken Korea

article

i n f o

Article history: Received March 1, 2011 Received in revised form September 8, 2011 Accepted September 8, 2011 Available online September 28, 2011 Keywords: PCB inventory Emission factor South Korea

a b s t r a c t The content of polychlorinated biphenyls (PCBs) was analyzed in samples collected from facilities associated with products or wastes containing PCBs in South Korea. Mean concentrations of atmospheric 209 PCBs were 7420 (37.0-104.048) pg m−3 and 16.8 (ND-34.2) fg WHO-TEQ m−3 in indoor air samples. and 1670 (106–13.382) pg m−3 and 5.64 (ND–36.0) fg WHO-TEQ m−3 in outdoor air samples. The highest levels were observed in indoor air samples from disposal facilities (7336–104,048 pg m−3), followed by manufacturing (330–25,057 pg m−3), recycling and storage facilities, indicating that PCB emissions from products containing PCBs and wastes remain very high and the facilities associated with them can be a significant source of atmospheric PCBs. Principal component analysis of PCB profiles showed that the homologous patterns of PCBs in outdoor and indoor air samples collected from facilities are similar to those of threshold air samples and commercial PCB products, e.g. Aroclor 1016, 1221, 1232 and 1242. Assessment of PCB mass balance in the plant, separation and washing of transformers contaminated with PCB, showed that approx. 0.0022% of the total PCB processed at this facility was released to the atmosphere, and the majority was transferred to the atmosphere. waste oil for disposal by burning or chemical methods. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Polychlorinated biphenyls (PCBs) are classified and regulated as one of the 12 persistent organic pollutants (POPs) according to the Stockholm Convention on POPs [1]. Sources of PCBs can be divided into two broad categories: intentionally produced chemicals in the chemical industry and by-products unintentionally synthesized de novo during thermal processes [2,3]. Global production and consumption of PCBs for industrial purposes is relatively well established. PCBs were mainly produced commercially from 1929 until the early 1970s. During this period, the total world production of PCBs was estimated at about 1.3 million tons [4]. Commercial PCBs are known by a number of trade names, including Aroclor (USA, UK), Kanechlor (Japan), Sovol (Russia), Chlophen (Germany, Poland) and Phenoclor (France) [5,6].

∗ Correspondence to: School of Environmental Science and Engineering, Pohang University of Science and Technology (POSTECH), San 31, Hyoja-dong, Namgu, Pohang 790-784, Republika Koreja. Tel.: +82 54 279 2281; Fax: +82 54 279 8299. E-mail address:[email protected](Y.-S. Chang). 1 Present address: Changbai Mountain Key Laboratory of Natural Resources and Functional Molecular (Yanbian University), Ministry of Education, 133002, Park Road 977, Yanji, Jilin Province, China. 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.030

Industrial mixtures of PCBs have never been produced in South Korea, and their use in electronic equipment has been banned since 1979, and their import and use were completely banned in 1996. Kim et al. [7] reported that the ambient air in Korea was more affected by the combustion process than in Japan, and also the contribution of commercial PCB products was relatively small. PCB levels in iron and steel complexes in South Korea have been reported to be higher than in residential areas, indicating that iron and steel complexes are likely to be an important source of PCBs [8]. However, PCB emission caused by de novo synthesis is not considered to be a major contributor to the global historical PCB mass balance [9]. The relative importance of atmospheric emissions from different source classes is not well known with considerable uncertainty [10]. Jamshidi et al. [11] reported that the main contemporary source of PCBs in UK urban centers is indoor air ventilation rather than soil evaporation. According to Korean law, waste containing PCBs (>0.0001 mg kg−1 in solids or >0.01 mg kg−1 in liquids) is considered "waste containing PCBs" that must be treated by special methods [12]. Recycling of waste containing PCBs is limited only to waste containing less than 2 mg kg−1 PCBs. In 2007, the amount of PCB-containing waste produced in South Korea contaminated with >2 mg kg−1 PCB was 2543 tons [13]. Therefore, the emission of PCBs from products containing PCBs i

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Fig. 1. PCB levels in air samples collected by HVAS from facilities associated with products or wastes containing PCBs. Cross bars represent mean values, and vertical lines represent maximum and minimum concentrations.

Waste remains very high even 30 years after PCB production ceases. Emissions inventories are crucial for the identification, assessment and prioritization of reasonable control strategies at the regional or global level [1,14]. To understand and predict long-range transport properties and fate of these substances in the environment, quantitative information on their releases to the atmosphere is also considered crucial [15,16]. Hosomi et al. evaluated PCB leaching from PCB-containing ballast in a fluorescent tube [17]. However, few studies have examined PCB contamination in facilities associated with products and wastes containing PCBs rather than unintentional sources such as incinerators. no such assessment has been conducted in South Korea. In this study, we investigated the atmospheric levels and distribution of PCBs in facilities associated with products and wastes containing PCBs. These facilities include manufacturing, utilization, recycling, storage and disposal throughout South Korea. We also estimated PCB emission factors and mass balance in a PCB disposal facility. PCB emission caused by de novo synthesis was not considered in this study. This is the first study to investigate PCB emissions from facilities associated with PCB-containing products and wastes in South Korea, providing valuable data in planning the comprehensive management and eventual disposal of PCB-containing products and wastes. 2. Examination and method 2.1. Sampling Air samples were collected from 44 sites (9 manufacturing, 8 in-use, 14 storage, 10 recycling, 1 disposal, and 2 borders) associated with products or waste containing PCBs throughout South Korea from October 2007 to July 2008 (Fig. S1). Sampling-specific data are presented in Table S1. Samples were collected using high volume air sampling (HVAS, DHA-1000S, SIBATA). A glass fiber filter (GFF) and two consecutive polyurethane foam (PUF) plugs were used to collect suspended particles or PCBs in the vapor phase. Before sampling, the GFFs were baked at 450 ◦ C for 12 h, and the PUF discs were Soxhlet-extracted for 16 h with acetone, then for 16 h with dichloromethane, and then dried in a vacuum desiccator for 24 h. A total of 20 outdoor air samples and 39 indoor air samples were collected during 24 hours with a flow rate of 700 l/min. It is important to note that

The sizes of the rooms of some buildings in use were smaller than the collected volume of air (1000 m3). Therefore, PCB concentrations could be underestimated by dilution effects in these small plants. Outdoor air samples were collected at a distance of 5 meters from the room (or rooms). The PCB concentration threshold was measured at the locations located on the borders (500-800 m) of the facilities. An additional 89 soil samples were collected in 37 facilities by dusting the floor with glass wool washed with hexane. At each site, 1-3 samples were collected from the bottom, depending on the size of the plant. For the mass balance case study, several samples of end products such as copper, silicon steel sheet, waste paper, used oil, etc. were collected from a PCB-containing waste disassembly and cleaning facility. 2.2. Analytical methods In the laboratory, the samples were processed, extracted and analyzed according to the methods prescribed by the US EPA method 1668A [18]. Briefly, samples were labeled with an internal standard containing 27 congeneric 13-labeled PCBs (1, 3, 4, 15, 19, 37, 54, 77, 81, 104, 105, 114, 118, 123, 156, 156, 156 , 156, 167, 169, 188, 189, 202, 205, 206, 208 and 209) (Wellington, 1668-LCS), then Soxhlet extracted for 24 hours using toluene. The extracts were then washed with concentrated H 2 SO 4 and then with saturated hexane H 2 O. Sample purification was performed using multi-layered silica and florisil columns. The eluent was reduced to 0.5 mL by rotary evaporation and a gentle stream of N2 gas. Finally, the extracts were transferred to GC vials and 13 C-labeled PCBs (9, 28, 52, 101, 111, 138, 178, and 194) were added as recovery standards. PCB content was analyzed using an Agilent Hewlett-Packard 6890 gas chromatograph/Jeol JMS-700T high-resolution mass spectrometer (GC/HRMS) with a DB-5MS column (J&W Scientiﬁc, 60 m length, 0.25 mm ID, 0.25 mm with a thickness of 0.25 mm). The instrument was operated using He as carrier gas at a constant flow rate of 1 ml min-1. The temperature program for the GC oven was as follows: the temperature was maintained at an initial value of 110 ◦ C for 2 min, then increased to 40–200 ◦ C min−1 and maintained for 3 min, then increased to 2–230 ◦ C min− 1, then increased to 7-300 ◦ C min−1 and held for 7 min. A 1 lL sample was injected at a temperature between 280 and 300 ◦ C to analyze the PCB content. The GC/HRMS was operated under positive EI conditions (38 eV) with a resolution of 10,000. Data were collected in selected ion tracking mode (SIM).

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Fig. 2. PCB levels in bottom samples. Cross bars represent mean values, and vertical lines represent maximum and minimum concentrations.

Peak assignment was performed to quantify 209 PCB congeners, but typically only 120 PCB congeners were detected. Bottom samples were analyzed using a GC with an electron capture detector (HP 6890, Agilent) according to the official Korean waste method [19].

3. Results and discussion 3.1. PCB Levels PCB levels detected were generally lower in outdoor air samples than in indoor air samples, although the range was very wide (Table 1, Figure 1). The average concentration of PCBs (209 PCBs) in outdoor air samples was 1670 pg m−3 (5.64 fg WHO-TEQ m−3) and varied from 106 pg m−3 at the PCB-containing waste repository to 13 400 pg m−3 3 in a PCB disposal facility (separation and cleaning). These PCB concentrations in outdoor air samples were consistent with PCB levels in South Korean ambient air in a previous study [7] and comparable to those in global urban areas (average: 1700 pg m−3) [20]. The average PCB concentration in indoor air samples was 7420 pg m−3 (16.8 fg WHO-TEQ m−3) and ranged from 37 pg m−3 in a closed transformer room (containing 46.6 tons of contaminated transformer oil with 0.15 mg kg−1 PCB) to 104.048 pg m−3 in a PCB disposal facility. PCB levels in indoor and outdoor air samples were highest in the disposal facility, followed by the manufacturing facility, recycling facility, storage area, and application area. A high concentration of PCBs in a landfill may be a consequence of its evaporation

2.3. Quality Assurance/Quality Control Several steps were taken to obtain data that would allow the accuracy and reliability of the data to be assessed. Analytical blanks were included in the amount of one per 10 samples. All data are blank corrected. Average recoveries of 27 PCB-labeled congeners 13 ranged from 25 to 93% (Table S2), which met the criteria (25–150%) recommended by US EPA Method 1668A. Recovery statistics are listed in Table S6. The detection limit of the method was calculated as 3 times the standard deviation of seven blank samples (Table S3). The criteria for analyte quantification were as follows: retention time within 2 s of that of the standard, isotopic ratio within 20% of that standard, and signal-to-noise ratio ≥3.

Table 1. Levels of PCBs (209 PCBs) in indoor and outdoor air samples from different locations using HVAS. Type of placement

Type of sample

Average PCB concentration 3

Marginal production In use (inside) In use (inside) In use (outside) Storage (inside) Storage (inside) Storage (outside) Recycling Recycling Disposal Disposal Historical industry Housing in use capacitor (contains PCB) Industrial area Urban area Historical

Outdoors (n = 2) Indoors (n = 9) Indoors (n = 5) Outdoors (n = 3) Outdoors (n = 3) Indoors (n = 11) Outdoors (n = 9) On outdoors Indoors (n = 1) Indoors (n = 10) Outdoors (n = 1) Indoors (n = 4) Outdoors (n = 2) Outdoors

Internal External (n = 3) External (n = 3) External

Literature −3

(str/m )

(fg WHO-TEQ m

153 (146-161) 8722 (330-25.057) 788 (37-2273) 815 (199-1159) 1017 (671-1561) 1469 (353-5404) 790 (37-720) 910 (37-720) –17.710 ) 1667 (1667–1667) 33.692 (7336–104.048) 8257 (3131–13.382) 180–280 2080–5820 240–28.000–28.00

1.44 (0.029–2.85) 11.50 (0.073–31.0) 4.79 (0.157–12.3) 8.20 (ND–24.6) 8.76 (ND–14.9) 0,72 (ND–3,74) 6,833 (ND 74) 6,83–6 ,81 41,3) ΝΔ 114 (ND–342) 2,31 (1,70–2,91)

21-27 19-46 0,6 (1-1,9)

) This study This study This study This study This study This study This study This study This study This study This study This study This study Kim et al. [7]

Hosomi [17] Martinez et al. [29] Menichini i south. [30]

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πιο πτητικά συγγενή PCB κατά την αποσυναρμολόγηση του μετασχηματιστή και το πλύσιμο με διαλύτες. Τα επίπεδα PCB διερευνήθηκαν επίσης σε 89 δείγματα πυθμένα που συλλέχθηκαν από 36 εγκαταστάσεις (Εικ. 2). Τα PCB ανιχνεύθηκαν στο 80% περίπου των δειγμάτων πυθμένα. Οι συγκεντρώσεις PCB κυμαίνονταν από ND έως 342 ng cm−2 και το υψηλότερο επίπεδο βρέθηκε σε δείγματα από μια τοποθεσία παραγωγής λαδιού μετασχηματιστή (ND–342 ng cm−2), ακολουθούμενη από μια εγκατάσταση διάθεσης (4–152 ng cm−2). Οι συγκεντρώσεις PCB 26.000–110.000 pg m−3 έχουν παρατηρηθεί στον εσωτερικό αέρα ενός γραφείου όπου είχαν χρησιμοποιηθεί λαμπτήρες φθορισμού με έρμα που περιέχει PCB [17]. οι ρυθμοί εξαέρωσης PCB από αυτό το έρμα ήταν εξαρτώμενοι από τη θερμοκρασία και η σύνθεση PCB του αερίου εκπομπής ήταν παρόμοια με αυτή που παρατηρήθηκε στα δείγματα έρματος που συλλέχθηκαν. Τα δείγματά μας δεν συλλέχθηκαν για να ποσοτικοποιηθεί η επίδραση της θερμοκρασίας στα επίπεδα PCB σε δείγματα αέρα εσωτερικού χώρου. Ωστόσο, οι υψηλότερες συγκεντρώσεις PCB παρατηρήθηκαν σε δείγματα που συλλέχθηκαν σε μια εγκατάσταση απόρριψης κατά τη διάρκεια του καλοκαιριού (104.000 pg m−3 τον Ιούλιο έναντι 7340 pg m−3 τον Οκτώβριο), ακολουθούμενη από μια μονάδα παραγωγής (25.100 pg m−3). Σε ένα μεγάλο αστικό συγκρότημα του Ηνωμένου Βασιλείου, η κύρια σύγχρονη πηγή PCB έχει αναφερθεί ότι δεν είναι η εξάτμιση από το έδαφος αλλά ο αερισμός του εσωτερικού αέρα. Οι υπάρχουσες κατασκευές, ειδικά τα παλαιότερα κτίρια στα οποία είχαν χρησιμοποιηθεί PCB στο παρελθόν, ήταν η κύρια πηγή PCB στον εξωτερικό αέρα [11]. Γενικά, οι αστικές περιοχές είναι πιο μολυσμένες από PCB από τις αγροτικές περιοχές [21]. Με βάση τα δεδομένα μας και τις προηγούμενες μελέτες μας, φαίνεται ότι τα PCB που εξατμίζονται από προϊόντα ή απόβλητα που περιέχουν PCB αποτελούν σημαντικές πηγές PCB στον ατμοσφαιρικό αέρα στη Νότια Κορέα. Μελλοντικές μειώσεις στις συγκεντρώσεις PCB στον εξωτερικό αέρα και τελικά στην έκθεση του ανθρώπου μπορεί να επιτευχθούν καλύτερα με ενέργειες σε αυτές τις υπόλοιπες πηγές PCB από προϊόντα και απόβλητα που περιέχουν PCB. Στα δείγματα αέρα μας, η μέση συμβολή των PCB αέριας φάσης στο σύνολο των PCB ήταν περίπου 96%, κάτι που ήταν σύμφωνο με την προηγούμενη μελέτη 24 ομοειδών PCB στη Νότια Κορέα [22]. Ο διαχωρισμός με αέρια σωματίδια PCB σε δείγματα αέρα από κάθε τύπο εγκαταστάσεων που σχετίζονται με προϊόντα ή απόβλητα που περιέχουν PCB έδειξε παρόμοια μοτίβα (Εικ. S5). Η συνεισφορά της αέριας φάσης στο σύνολο των PCB μειώθηκε από 89 (μονο-CB) σε 24% (deca-CB) σε δείγματα αέρα εσωτερικού χώρου και από 92 (μονο-CB) σε 22% (deca-CB) σε δείγματα εξωτερικού αέρα (Εικ. S6). Αυτά τα αποτελέσματα υποδεικνύουν ότι τα PCB στον αέρα υπάρχουν κυρίως στην αέρια φάση και ότι η συμβολή των ομοειδών PCB στην αέρια φάση μειώνεται καθώς τα ομοειδή χλωριώνονται περισσότερο (δηλαδή, λιγότερο πτητικά). 3.2. Μοτίβα ομολογίας Τα ομόλογα σχέδια των PCB στα δείγματα αέρα ήταν παρόμοια σε όλες τις τοποθεσίες δειγματοληψίας (Εικ. S2). Σε όλα τα δείγματα αέρα, τα κυρίαρχα ομόλογα PCB που βρέθηκαν ήταν PCB χαμηλής χλωρίωσης, όπως μονο-, δι και τρι-PCB, τα οποία αντιπροσώπευαν περίπου το 14%, 35% και 33% των συνολικών PCB σε δείγματα αέρα εσωτερικού χώρου και περίπου 15%, 41% , και 30% των συνολικών PCB σε δείγματα εξωτερικού αέρα, κατά μέσο όρο, αντίστοιχα (Εικ. S3). Πολλές πηγές PCB μπορούν να επηρεάσουν τα ατμοσφαιρικά επίπεδα PCB, συμπεριλαμβανομένων των αποτεφρωτηρίων, των βιομηχανικών θερμικών διεργασιών και των προϊόντων και αποβλήτων που περιέχουν PCB [23]. Τα ομόλογα πρότυπα μπορούν να παρέχουν ενδείξεις για το πού και πώς προήλθαν αυτές οι ουσίες [24]. Τα PCB στον εξωτερικό αέρα σε αυτή τη μελέτη προφανώς επηρεάστηκαν από τα PCB του εσωτερικού αέρα λόγω των υψηλότερων επιπέδων PCB στο εσωτερικό περιβάλλον. Για περαιτέρω προσδιορισμό της πηγής, η ανάλυση κύριου συστατικού (PCA) των δεδομένων διεξήχθη χρησιμοποιώντας λογισμικό SPSS 12.0 (SPSS, Inc.) και τα ομόλογα μοτίβα δειγμάτων αέρα από αυτήν τη μελέτη συγκρίθηκαν με αυτά των εμπορικών μιγμάτων του Aroclor 1016, 1221, 1232, 1242, 1248, 1254, 1260, 1262 και 1268 από άλλες μελέτες [6,25]. Οι συνολικές συγκεντρώσεις κάθε ομολόγου PCB (δηλ. 1 Cl, 2 Cls, .. .., 10 Cls) χρησιμοποιήθηκαν για ανάλυση PCA. Τα δεδομένα PCB κανονικοποιήθηκαν με διαίρεση με τις συνολικές συγκεντρώσεις PCB για κάθε δείγμα, παράγοντας δεδομένα που κυμαίνονται από 0 έως 1. Τέλος, αυτές οι κανονικοποιημένες συνθέσεις PCB χρησιμοποιήθηκαν ως δεδομένα εισόδου για PCA. Σαν άποτέλεσμα,

Τα PC1 και PC2 αντιπροσώπευαν το 60% της συνολικής διακύμανσης (Εικ. 3). Στο διάγραμμα φόρτωσης, οι μεταβλητές είναι καλά ομαδοποιημένες με βάση τον αριθμό του χλωρίου. Τα ομόλογα σχέδια των PCB σε δείγματα αέρα από διάφορες τοποθεσίες σε αυτή τη μελέτη ήταν παρόμοια με εμπορικά μείγματα όπως τα Aroclor 1016, 1221, 1232 και 1242, υποδηλώνοντας ότι τα ομόλογα σχέδια πολλών δειγμάτων αέρα επηρεάστηκαν ταυτόχρονα από αυτά τα εμπορικά μείγματα. Είναι συνεπές με μια άλλη προηγούμενη μελέτη ότι τα ομόλογα πρότυπα των PCB που βρέθηκαν σε ιζήματα στη Νότια Κορέα έδειξαν ότι οι πηγές τους ήταν εμπορικά μείγματα όπως Aroclor 1016, 1242, 1254 και 1260 ή αντίστοιχα προϊόντα Kanechlor [26]. Το PCA χρησιμοποιήθηκε επίσης για τη σύγκριση ομόλογων μοτίβων δειγμάτων αέρα σε αυτή τη μελέτη με εκείνα του ατμοσφαιρικού εδάφους [27], των καυσαερίων αποτέφρωσης και των δειγμάτων καυσαερίων εργοστασίων τσιμέντου από άλλες έρευνες [6,28]. Ως αποτέλεσμα, τα PC1 και PC2 αντιπροσώπευαν το 90% της συνολικής διακύμανσης (Εικ. 4). Τα ομόλογα μοτίβα των PCB στα δείγματα αέρα μας από διάφορες τοποθεσίες στη Νότια Κορέα ήταν διαφορετικά από εκείνα του ατμοσφαιρικού εδάφους, των καυσαερίων αποτέφρωσης και των δειγμάτων καυσαερίων εργοστασίων τσιμέντου, υποδηλώνοντας ότι υπάρχουν και άλλες σημαντικές πηγές. Ωστόσο, όλα τα δείγματα αέρα, συμπεριλαμβανομένων των δειγμάτων οριακού αέρα είχαν παρόμοια ομόλογα μοτίβα με τα γενικά δείγματα αέρα περιβάλλοντος (Κορέα, n = 15, Ιαπωνία, n = 11) [7], δείγματα αέρα εσωτερικού χώρου από μια εγκατάσταση διάθεσης (Ιαπωνία, n = 5 ), και δείγματα αέρα εσωτερικού χώρου ενός δωματίου όπου χρησιμοποιήθηκε στεγανωτικό που περιέχει PCB (Ιαπωνία, n = 3) [6] (Εικ. S4). Οι Kim et al. [7] ανέφερε ότι τα επίπεδα PCB στον ατμοσφαιρικό αέρα της Νότιας Κορέας επηρεάστηκαν περισσότερο από τις διαδικασίες καύσης από ό,τι στην Ιαπωνία, και επίσης ότι η συμβολή των εμπορικών προϊόντων PCB ήταν σχετικά μικρή. Ωστόσο, τα αποτελέσματά μας υποδηλώνουν έντονα ότι ο ατμοσφαιρικός αέρας στη Νότια Κορέα είναι μολυσμένος από μείγματα εμπορικών προϊόντων Aroclor με διάφορες περιεκτικότητες σε χλώριο, ιδιαίτερα από μείγματα χαμηλής χλωρίωσης. 3.3. Μια μελέτη περίπτωσης ισοζυγίου μάζας PCB Για να αξιολογηθεί η ισορροπία μάζας PCB, επιλέχθηκε μια εγκατάσταση απόρριψης PCB, όπου τα PCB στον εσωτερικό αέρα έδειξαν το υψηλότερο επίπεδο. Μια σειρά δειγμάτων αέρα, πυθμένα και τελικού προϊόντος συλλέχθηκαν από κάθε διαδικασία αποσυναρμολόγησης και καθαρισμού. Αυτή η εγκατάσταση επεξεργάζεται κυρίως απόβλητα μετασχηματιστών που περιέχουν λάδι μετασχηματιστή μολυσμένο με PCB (>2 mg kg−1 ). Οι κύριες διαδικασίες είναι η αφαίρεση του λαδιού του μετασχηματιστή, η αποσυναρμολόγηση των εξωτερικών περιβλημάτων του μετασχηματιστή, η πρώτη εξαγωγή με τολουόλιο, η αποσυναρμολόγηση των εσωτερικών συγκροτημάτων του μετασχηματιστή και η δεύτερη εξαγωγή με τολουόλιο. Τα τελικά προϊόντα ανακυκλώνονται (μέταλλα) ή περνούν για περαιτέρω διάθεση (απορρημένα λιπαντικά). Οι μέσες συγκεντρώσεις PCB μετρήθηκαν ως 14.560 pg m−3 σε δείγματα αέρα εσωτερικού χώρου και 8130 pg m−3 σε δείγματα εξωτερικού αέρα (Εικ. 5). Δεδομένου ότι όλες οι δειγματοληψίες PCB διενεργήθηκαν μόνο την φθινοπωρινή περίοδο, δεν αντικατοπτρίστηκε καμία εποχική διακύμανση των PCB στον εξωτερικό αέρα. Αυτός είναι ένας από τους περιορισμούς στην παρούσα μελέτη. Τα περιορισμένα δεδομένα παρακολούθησης πεδίου χρήζουν περαιτέρω μελέτης στο μέλλον. Εάν οι συγκεντρώσεις των PCB στον εσωτερικό αέρα είναι σχετικά σταθερές κατά την περίοδο λειτουργίας, η ισορροπία PCB αέρα μεταξύ εσωτερικού και εξωτερικού αέρα μπορεί να περιγραφεί από την ακόλουθη εξίσωση. (1): V

dC1 = Ea Q − VR (C1 − C0) = 0 dt

(1)

and the PCB air emission factor Ea in this facility can be calculated using Eq. (2), Ea =

(C1 − C0)VR Q

(2)

where V is the volume (m3), C1 is the PCB concentration in the indoor air (ng m−3), C0 is the PCB concentration in the outdoor air (ng m−3), t is the time, Q is the amount of PCB processed in the facility , and R is the natural air exchange rate at which the outside air replaces all the inside air. In general, the natural air exchange of a concrete building is 7-24 days−1 [17],

G.-Z. Jin i sur. / Journal of Hazardous Materials 196 (2011) 295–301

Fig. 3. Comparison of homology plots of air samples from this study with commercial PCB mixtures.

Fig. 4. Comparison of homologous standards of air samples from this study with samples from other sources.

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Products and waste containing PCBs in South Korea. Total PCB concentrations ranged from 37.0 to 104,048 pg m−3 in indoor air samples and from 106 to 13,382 pg m−3 in outdoor air samples. Homologous patterns of PCBs in outdoor and indoor air samples collected from different facilities were similar to patterns of boundary air samples and commercial PCB mixtures Aroclor 1016, 1221, 1232 and 1242. These results suggest that PCB emissions during manufacturing, recycling, use and disposal of products and waste containing PCBs can be a significant source of PCBs in the atmosphere. As such, it provides valuable information in planning the comprehensive management and eventual disposal of products and waste containing PCBs. Fig. 5. Mass balance of PCBs in the PCB disposal facility.

Acknowledgments used in this study. Using data observed at this facility (Table S4), Ea was estimated to be 9.8 × 10-4 to 3.4 × 10-3 g-PCB gPCB-1 yr-1. This value is comparable or slightly higher than that published in a previous study [4]. Ea ranged from 1.58 × 10-5 to 2.56 × 10-2 g-PCB g-PCB-1 yr-1 for open use and storage sites and from 3.38 × 10-9 to 5, 22 × 10-4 g-PCB g-PCB −1 year−1 for indoor use and storage [9]. In a previous study, Breivik et al. [4] reported emission factors for only 22 PCB congeners, which have high uncertainty. Specific amounts of 22 congeners in Aroclor mixtures in South Korea are not available. However, a direct comparison of the PCB air emission factors reported in this study with those reported above may be reasonable because 22 congeners were the dominant PCB congeners and both air emission factors had the same unit. The uncertainty in the calculation of the air emission factor in this study mainly comes from the value of natural air exchange, variations of PCB concentrations in indoor and outdoor air samples, temperature, etc. (3): Eb =

Cb A Q

(3)

where Cb is the average PCB concentration in bottom samples, and A is the surface of the object. Using data measured at this facility, Eb was calculated for bottom samples to be 3.0 × 10-4 gPCB g-PCB-1 yr-1. The mass balance of PCBs at this landfill was calculated (Fig. 5). There are great uncertainties, because Breivik et al. [9], which mainly results from the value of natural air exchange, variation of PCB concentrations in air samples and temperature, etc. Of the total PCBs disposed of at this facility, approx. 0.0022% into the atmosphere, and 0.03% settles on the indoor floor as dust particles or transformer oil leaks. Meanwhile, most PCBs (98.7%) were transported as transformer oil spills for later disposal by incineration or chemical treatment. If this plant were to operate at a maximum capacity of 100 tons/week, the estimated maximum emission of 209 PCBs into the air would be 13 g/year. This is much lower than the previous estimate, where the estimated air emissions of PCBs in South Korea were 199 kg (for 22 similar units of PCBs, the intermediate scenario, the maximum is several hundred times higher than the minimum scenario) in the reference year 2008 [4 ]. Although the production, import and use of PCBs have been banned in South Korea since 1999, the amount of products containing PCBs in use is still huge. In South Korea, the amount of PCB-containing waste, which was contaminated with >2 mg kg−1 PCB, was 2543 tons in 2007 [13]. Therefore, atmospheric emission of PCBs from products and wastes containing PCBs may be a significant source for some time until they are completely eliminated. 4. Conclusion In this study, we investigated PCBs from different types of plants and calculated the mass balance of PCBs in the plant in relation to

This work was supported by a grant from the National Research Foundation of Korea (NRF) funded by the Korean government (MEST) (No. 2011-1128723) and partially supported by the Korea Institute of Science and Technology (KIST) as an institutional program ( 2E22173 ). Appendix A. Supplementary data Supplementary data related to this article can be found online at doi:10.1016/j.jhazmat.2011.09.030. References [1] UNEP, Stockholm Convention on Persistent Organic Pollutants (POPs), chemicals, United Nations Environment Programme, 2001 [2] S. Sakai, M. Hiraoka, N. Takeda, K. Shiozaki, Formation and emission of nonortho CBs and mono-ortho CBs in municipal waste incineration, Chemosphere 29 (1994) 1979-1986. [3] B. Wyrzykowska, N. Hanari, A. Orlikowska, N. Yamashita, J. Falandysz, Dioxin composition of bottom ash from household combustion in Poland and their possible correlations with the pollution status of agricultural land and pine needles, Chemistry 76 (2009) 255-263. [4] K. Breivik, A. Sweetman, J.M. Pacyna, K.C. Jones, Towards a global historical emission inventory for selected similar PCBs - a mass balance approach. 3. Update, Sci. The whole environment. 377 (2007) 296-307. [5] S.K. Shin, T.S. Kim, Levels of polychlorinated biphenyls (PCBs) in transformer oils from Korea, J. Hazard. Mater. 137 (2006) 1514-1522. [6] Y. Ishikawa, Y. Noma, Y. Mori, S.-i. Sakai, PCB congener profiles and a proposed new set of congener markers, Chemosphere 67 (2007) 1838-1851. [7] K.S. Kim, B.-J. Song, J.-G. Kim, K.-K. Kim, Investigation of pollution levels and sources of polychlorinated biphenyl (PCB) in ambient air in Korea and Japan, J. KSEE 27 (2005) 170-176. [8] S.-D. Choi, S.-Y. Baek, Y.-S. Chang, Passive air sampling of persistent organic pollutants in Korea, Toxicol. Surround. Health Sci. 1 (2009) 75-82. [9] K. Breivik, A. Sweetman, J.M. Pacyna, K.C. Jones, Toward a global historical emissions inventory for selected PCB congeners—a mass balance approach: 2. Emissions, Sci. The whole environment. 290 (2002) 199-224. [10] K. Breivik, R. Alcock, Y.F. Lee, R.E. Bailey, H. Fiedler, J.M. Pacyna, Primary sources of selected POPs: regional and global emission inventories, Environ. Pollution. 128 (2004) 3-16. [11] A. Jamshidi, S. Hunter, S. Hazrati, S. Harrad, Concentrations and chiral signatures of polychlorinated biphenyls in outdoor and indoor air and soil in a large urban area of the United Kingdom, Environ. Sci. Technol. 41 (2007) 2153-2158. [12] Korea, Special Law on the Management of Persistent Organic Pollutants, Ministry of Environment, 2008. [13] Korea, Status of Generation and Disposal of Specific Waste, Ministry of Environment, 2008. [14] UN/ECE, 1998 Aarhus Protocol on POPs, United Nations/Economic Council for Europe, 1998. www.unece.org/env/lrtap/pops h1.htm. [15] K. Breivik, B. Bjerkeng, F. Wania, A. Helland, J. Magnusson, Modeling the fate of polychlorinated biphenyls in Inner Oslofjord, Norway, Environ. Toxicol. Chem. 23 (2004) 2386-2395. [16] H. Hung, C.L. Sum, F. Wania, P. Blanchard, K. Brice, Measurement and simulation of atmospheric concentration trends of polychlorinated biphenyls in the Northern Hemisphere, Atmos. Surround. 39 (2005) 6502-6512. [17] M. Hosomi, Volatilization of PCBs from PCB-containing ballasts in fluorescent lamps and indoor PCB pollution: ood of PCBs, J. Japan Assoc. The smell of the environment. 36 (2005) 323-330. [18] USEPA, Method 1668, Revision A: Congeners of Chlorinated Biphenyls in Water, Soil, Sediment, and Tissues by HRGC/HRMS, 1999. [19] Korea, Official Method for the Analysis of PCBs in Waste Samples, Ministry of the Environment 2005.

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Journal of Hazardous Materials 196 (2011) 263-269

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Phenanthren og pyren radikale nedbrydningsgradienter i mangrove (Kandelia candel (L.) Druce) sediment Haoliang Lu a,b , Yong Zhang a,∗ , Beibei Liu a , Jingchun Liu b , Juan Ye b , Chongling Yan b a b

State Key Laboratory of Marine Environmental Science (Xiamen University), Environmental Research Center, Xiamen University, Xiamen 361005, Fujian Province, PRC Ministry of Education Key Laboratory of Eastern and Wetland Ecosystems and School of Natural Sciences, Xiamen University, Xiamen 361005, Φυτζιάν, LΚίνα

article

i n f o

Article history: Received January 21, 2011 Received in revised form September 7, 2011 Accepted September 7, 2011 Available online September 14, 2011 Keywords: Rhizodegradation Phenanthrene Pyrene Mangrove Kandelia (L.) Druce

a b s t r a c t A greenhouse experiment was conducted to evaluate the rate of degradation of phenanthrene (Ph, 10 mg kg−1) and pyrene (Py, 10 mg kg−1) in the rhizosphere of the mangrove candelium Kandelia (L.) Druce. The Rhizosphere model system was created using an improvised lamellar root box that was divided into eight separate compartments at different distances from the root surface. After 60 days of plant growth, the presence of the plant significantly increased the diffusion of Ph (47.7%) and Py (37.6%) from the contaminated sediment. Higher PAH degradation rates were observed 3 mm from the root zone (56.8% Ph and 47.7% Py). The degradation gradient followed the order: near rhizosphere > root space > far rhizosphere soil zone for both pollutants where mangrove was grown. The contribution of direct plant uptake and accumulation of Ph and Py was very low compared to plant-facilitated diffusion. In contrast, plant-promoted biodegradation was the dominant contributor to enhanced recovery. Correlation analysis shows a negative correlation between biological activities (microbial biomass carbon activity, dehydrogenase, urease and phosphatase activities) and residual concentrations of Ph and Py in planted soils. Our results suggest that the mangrove rhizosphere was effective in promoting the depletion of aromatic hydrocarbons in polluted sediments. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Polycyclic aromatic hydrocarbons (PAHs) are ubiquitous pollutants that remain in the environment. Anthropogenic input of PAHs from oil spills, shipping, urban runoff and emissions from combustion and industrial processes has caused significant accumulation of PAHs in coastal mangrove swamps, especially those near urban centers and industrial cities [1,2]. Phytoremediation of PAHs is a promising alternative approach to sediment remediation due to its cost-effectiveness, practicality and environmental acceptability [1,3]. There are several branches of phytoremediation identified by USEPA (2000), including phytoextraction, rhizofiltration, phytoptosis treatment, rhizo- and phytodegradation, and phytostabilization. Rhizodegradation refers to the microbial breakdown of organic pollutants in soil and sediments in the root zone (rhizosphere). This process uses the natural ability of plants to manipulate the biological, chemical and physical properties of the rhizosphere to reduce the concentrations of organic pollutants in soil and sediments [4,5]. In sediment, rhizoremediation has been proposed as the main mechanism responsible for PAHs

∗ Corresponding author. Phone: +86 592 2188685; fax: +86 592 2184977. E-mail address:[email protected](Y. Zhang). 0304-3894/$ – see cover page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.031

degradation in plant-assisted restoration efforts [6,7]. In that case, the roots contribute to the spread of hydrocarbon pollutants by increasing the number of microbes, improving the physical and chemical conditions of the soil, increasing the exudate and wetting of the roots, and adsorbing pollutants in the rhizosphere. Mangrove ecosystems, important intertidal wetlands along the coasts of tropical and subtropical regions, are closely linked to industrial activities and exposed to pollution [8,9]. Mangroves may contribute to the diffusion of organic pollutants through increased microbial numbers, improved physical and chemical soil conditions, increased wetting and pollutant sorption in the rhizosphere, but the impact of each process is not clearly understood. A number of bacterial strains capable of degrading PAHs have been isolated from mangrove surface sediment, and the degradation of PAHs by these consortia and isolates in the culture medium and sediment pulp has been studied [10,11]. However, the question of how far the effect of the mangrove rhizosphere on PAH degradation may extend has never been addressed, but the preferential use of plants with fibrous root systems for rhizolysis suggests that it was quite limited. The production of protons, exudates and metabolites is released by the plant root into the soil rhizosphere, which has led to significant

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Οι διαφορές μεταξύ ριζόσφαιρας και μη ριζόσφαιρας στις ιδιότητες του εδάφους έχουν αναθεωρηθεί σε προηγούμενες εκθέσεις [12,13]. Η ριζόσφαιρα, ένα στρώμα εδάφους που περιβάλλει τις ρίζες των φυτών, ήταν δύσκολο να γίνει φυσική δειγματοληψία και χειρισμός με ακρίβεια. Το έδαφος της ριζόσφαιρας διαχωρίστηκε συνήθως από τη ρίζα του φυτού με ήπια ανακίνηση. Σε μια προσπάθεια να ξεπεραστούν ορισμένα από αυτά τα προβλήματα, σχεδιάστηκε ένα ριζόκουτο όπου το έδαφος κοντά στις ρίζες επέτρεψε τη συγκομιδή λεπτότερων διαδοχικών τμημάτων (1–5 mm και >5 mm) εδάφους ριζόσφαιρας στο εργαστήριο [14] . Πρόσφατα, η φυτοαποικοδόμηση ιζημάτων μολυσμένων με PAH με χρήση φυτών μαγγρόβια έχει αποτελέσει αντικείμενο αρκετών μελετών [2,8]. Ωστόσο, περιορισμένος από τεχνικές δειγματοληψίας εδάφους ριζόσφαιρας κοντά στις ρίζες, η εξαρτώμενη από την απόσταση μείωση της μικροκλίμακας των PAH στη διεπαφή ρίζας-εδάφους κατά μήκος της κλίσης της ριζόσφαιρας έχει μέχρι στιγμής σπάνια μελετηθεί. Ως εκ τούτου, υποθέσαμε ότι οι διαφορές με την απόσταση στα αποτελέσματα της ριζόσφαιρας θα συμπίπτουν με τις διαβαθμίσεις αποικοδόμησης των PAH. Ο στόχος αυτής της μελέτης ήταν επομένως η διερεύνηση της επίδρασης της ριζόσφαιρας στη διαδικασία απομάκρυνσης των PAH σε ένα ειδικά σχεδιασμένο ριζοσφαίριο που επέτρεπε τον διαχωρισμό του εδάφους ριζόσφαιρας (διαμέρισμα ρίζας) και του εδάφους που επηρεάστηκε από την έκκριση ριζών (διαμέρισμα χωρίς ρίζες). Η μελέτη επιχειρεί επίσης να αποκαλύψει την επίδραση της αυξανόμενης απόστασης στην αφαίρεση του ΠΑΥ. Οι PAH με 3 δακτυλίους Ph και οι PAH 4 δακτυλίων Py χρησιμοποιήθηκαν ως στόχοι PAH. Ως πρότυπο φυτό επιλέχθηκε Kandelia candel (L.) Druce (K. candel), ένα κοινό είδος κόκκινου μαγγρόβιου στην Κίνα. 2. Υλικά και μέθοδοι 2.1. Τα Chemicals Ph και Py με καθαρότητα 99,9% ελήφθησαν από τη Sigma–Aldrich Co. Ltd., UK. Όλες οι άλλες χημικές ουσίες που χρησιμοποιήθηκαν στη μελέτη ήταν αναλυτικής καθαρότητας. 2.2. Προετοιμασία ιζήματος με αιχμή PAH Μαζικά δείγματα επιφανειακών ιζημάτων συλλέχθηκαν από τον υγρότοπο μαγγροβίων Jiulong Estuary, PR China, και κοσκινίστηκαν μέσω κόσκινου 0,5 cm για να αφαιρεθούν τα χονδροειδή υπολείμματα, ομογενοποιήθηκαν και στη συνέχεια αποθηκεύτηκαν στους 4 ◦ C μέχρι τη χρήση. Οι φυσικές και χημικές ιδιότητες των ιζημάτων μετρήθηκαν στο εργαστήριο ως εξής: pH 6,63, περιεκτικότητα σε υγρασία 49,5%, συνολική οργανική περιεκτικότητα 2,1% και συνολική ποσότητα αζώτου 0,90 g kg−1 ξηρό ίζημα, συνολική ποσότητα φωσφορικών 0,62 g kg−1 ξηρό ικανότητα ανταλλαγής ιζημάτων και κατιόντων 15,8 cmol kg−1 . Οι PAH ανιχνεύθηκαν στα δείγματα ιζήματος με συγκέντρωση 19,3 ␮g kg−1 Ph και 24,56 ␮g kg−1 Py, αντίστοιχα. Το ίζημα λήφθηκε και εμπλουτίστηκε με PAH ως εξής: ένα μέρος του ιζήματος ζυγίστηκε με ακρίβεια στο δοχείο. Στη συνέχεια, ένας όγκος των PAHs διαλυμένων σε ακετόνη προστίθεται και αφήνεται να εξισορροπηθεί με τη μήτρα, αποθηκεύεται στο σκοτάδι και αφήνεται να στεγνώσει. Το ισοζύγιο μάζας χρησιμοποιείται για τον προσδιορισμό της εξάτμισης της ακετόνης. Η ακετόνη εξατμίστηκε 12 ώρες και το τμήμα του αιχμηρού ιζήματος αναμείχθηκε πρώτα με σχεδόν το 25% του συνολικού ιζήματος και στη συνέχεια για να αναμιχθεί με το υπόλοιπο 75% των υγρών ιζημάτων ακολουθούμενη από μηχανική ανάμειξη. Μετά από παλαίωση για 7 ημέρες, το ίζημα χρησιμοποιήθηκε για πείραμα rhizobox. Η ανιχνευθείσα συγκέντρωση του Ph και του Py ήταν 10 ± 0,5 και 10 ± 0,4 mg kg−1, αντίστοιχα σε ίζημα ηλικίας 7 ημερών. Δεν προστέθηκαν θρεπτικά συστατικά στο έδαφος κατά τη διάρκεια του πειράματος. 2.3. Πειραματικός σχεδιασμός Ένα εργαστηριακό rhizobox τροποποιημένο από την προηγούμενη μελέτη μας [15] (Εικ. 1) χρησιμοποιήθηκε για τη φύτευση κεριού K.. Η διάσταση του ριζόκουτου (Εικ. 1) ήταν 150 mm × 300 mm × 200 mm

Fig. 1. Schematic diagram of the rhizobox (modified according to Lu et al. [15]). S0: sediment for seedling growth. S1: rhizosphere; S2: close to the rhizosphere. S3: close to bulk soil. and S4: bulk bulk.

(length × width × height). The rhizobox was divided into five parts from the center to the left or right edge of the rhizobox, which was surrounded by a nylon fabric (400 mesh), i.e. the central plant growth zone (20 mm wide), the rhizosphere zone (1 mm wide), the nearby zone rhizosphere (2 mm wide), near-soil zone (10 mm wide) and soil mass zone (52 mm wide). In the rhizobox soil for seedling growth, the rhizosphere, near-rhizosphere, near-bulk, and bulk soil zones are designated as S0, S1, S2, S3, and S4, respectively. The design successfully prevents root hairs from penetrating adjacent soil zones, as well as keeping soil zones separate, while allowing transport of soil microfauna and root exudates between compartments. Approx. 12 kg of treated sediment was added to each root box, each treatment had three replicates. Five K. candelium seedlings were planted in the central zone of the box. The plants were grown in greenhouse conditions with natural light and relative humidity of 85%, the temperature varied from 26 to 32 ◦ C for 60 days. Sediment moisture content was adjusted to 100% water holding capacity by freshwater irrigation to reduce drainage and simulate anoxic water conditions. Rhizobox without plant was used as a control. Rhizoboxes were arranged in a random design in the greenhouse and their position was regularly rotated to ensure uniform conditions. Before harvesting, rice fields are not watered for 2 days. Harvesting involved sequential dissection of each root box, separation of layers in each rhizocyte soil zone, and removal of plants from the root compartment. Roots and shoots were manually separated from the soil, washed with deionized water and then dried with filter paper. Soil samples from different soil zones of each rhizocyte were homogenized separately before analysis. 2.4. Analyzes 2.4.1. Enzyme activity measurements The soil microbial biomass carbon (Cmic) was determined by the fumigation-extraction method with chloroform [16,17]. Sediment dehydrogenase activity was measured by the reduction of 2,3,5-triphenyltetrazolium chloride (TTC) to 1,3,5-triphenylformazan (TPF). Briefly, 5.0 g of lyophilized sediment sample was incubated for 24 hours at 37 ◦ C in 5.0 mL of TTC solution (5.0 g L-1 in 0.2 mol/L Tris-HCl buffer, pH 7.4). Two drops of concentrated H2SO4 were added immediately after incubation to stop the reaction. The sample was then mixed with 5.0 mL of toluene to extract TPF and shaken for 30 min at 250 rpm (25 ◦ C), followed by centrifugation

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στις 5000 rpm για 5 λεπτά, και η απορρόφηση στο εκχύλισμα μετρήθηκε στα 492 nm. Τέλος, η δραστηριότητα της αφυδρογονάσης του εδάφους υπολογίστηκε ως 1,00 ␮g TPF g−1 ξηρό ίζημα [18]. Χρησιμοποιήθηκε χρωματομετρική μέθοδος για τον προσδιορισμό της δράσης της ουρεάσης και της φωσφατάσης [19]. Οι ενζυμικές δραστηριότητες εκφρασμένες ως mg NH4 –N απελευθέρωσαν kg−1 ξηρό ίζημα στους 37 ◦ C και mg P απελευθέρωσαν kg−1 ξηρό ίζημα στους 37 ◦ C το καθένα για ουρεάση και φωσφατάση, αντίστοιχα. 2.4.2. Ανάλυση PAH Τα δείγματα ιζημάτων ξηράνθηκαν με ψύξη, δικτυώθηκαν και εκχυλίστηκαν με ένα επιταχυνόμενο σύστημα εξαγωγής μικροκυμάτων τροποποιημένο από τους Zhang et al. [20]. Εν συντομία, 10,00 g λυοφιλοποιημένου ιζήματος εκχυλίστηκαν με 50 mL διαλύτη μίγματος (η-εξάνιο/ακετόνη 1:1, ν/ν) χρησιμοποιώντας σύστημα εκχύλισης μικροκυμάτων (CEM Co., Matthews, NC, USA). Τα υποκατάστατα Ph-d10 και chrysene-d12 (Chy-d12) (Sigma–Aldrich, UK) προστέθηκαν στα δείγματα πριν από την εκχύλιση. Ενεργοποιημένος χαλκός (αναδευόμενος χαλκός με 5% διάλυμα ιωδιδίου/ακετόνης για περίπου 10 λεπτά) προστέθηκε στο εκχύλισμα για αποθείωση, και στη συνέχεια προ-συμπυκνώθηκε στα 2 mL με έναν περιστροφικό εξατμιστή (Buchi Vac V-800, Ελβετία). Τα συμπυκνωμένα εκχυλίσματα κλασματοποιήθηκαν με χρωματογραφία στήλης αλουμίνας/πυριτικής πηκτής (100–200 mesh) (40 cm × 1,5 cm εσωτερικά) συσκευασμένα από τον πυθμένα με υαλοβάμβακα, 10,00 g ουδέτερο οξείδιο του αργιλίου (100–200 mesh, ξηράνθηκαν στους 440 °C 4 ώρες), 18,00 g silica gel (100–200 mesh, ξηράνθηκε στους 170 ◦ C για 4 ώρες) και 2,00 g άνυδρου θειικού νατρίου. Οι αναλυτές στόχοι εκλούστηκαν από τη στήλη με 150 mL διαλύτη μίγματος η-εξάνιο/μεθυλενοχλωρίδιο (1:1, ν/ν). Αυτό το κλάσμα στη συνέχεια συμπυκνώθηκε στα 2 mL με εξάτμιση σε περιστροφικό κενό σε ένα λουτρό νερού στους 60 ◦ C και ο διαλύτης ανταλλάχθηκε σε η-εξάνιο. Το κλάσμα PAH συμπυκνώθηκε τελικά σε 1 mL υπό ήπιο ρεύμα αζώτου πριν από την ανάλυση GC/MS. Τα δείγματα των φυτών αλέστηκαν και ομογενοποιήθηκαν και εκχυλίστηκαν χρησιμοποιώντας την ίδια μέθοδο όπως και για το ίζημα. Οι συγκεντρώσεις των PAH στα εκχυλίσματα προσδιορίστηκαν με αέρια χρωματογραφία Hewlett-Packard 6890 εξοπλισμένη με ανιχνευτή φασματοσκοπίας μάζας (HP5975B). Η στήλη HP-5MS (Agilent Co., ΗΠΑ) είχε μήκος 30 m, με εσωτερική διάμετρο 0,25 mm και πάχος μεμβράνης 0,25 ␮m. Η θερμοκρασία αυξήθηκε από 60 ◦ C σε 150 ◦ C με ρυθμό 15 ◦ C min−1, αυξήθηκε σε 220 ◦ C στους 5 ◦ C min−1 και αυξήθηκε σε 300 ◦ C στους 10 ◦ C min−1, στη συνέχεια διατηρήθηκε στους 300 ◦ C για 5 λεπτά. Ως φέρον αέριο χρησιμοποιήθηκε ήλιο. Οι θερμοκρασίες του εγχυτήρα και του ανιχνευτή ήταν 280 ◦ C και 300 ◦ C, αντίστοιχα. Η ενέργεια κρούσης ηλεκτρονίων ήταν 70 eV και ο λόγος σάρωσης μάζας προς φορτίο (m/z) ήταν από 50 έως 400 amu. Επιλέχθηκε η επιλεγμένη λειτουργία ιόντων (SIM). Τα όρια ανίχνευσης που προέρχονται από αντίγραφα και διαδικαστικά κενά ήταν 2,2 και 1,6 ␮g kg−1 ξηρού βάρους για Ph και Py, αντίστοιχα. Όλα τα δεδομένα υπόκεινταν σε αυστηρές διαδικασίες ποιοτικού ελέγχου. Αιχμές μήτρας, διπλότυπα εργαστηριακών δειγμάτων και εργαστηριακά κενά υποβλήθηκαν σε επεξεργασία με κάθε παρτίδα δειγμάτων (10 δείγματα ανά παρτίδα) ως μέρος του εσωτερικού ελέγχου ποιότητας του εργαστηρίου. Οι μέσες ανακτήσεις δευτεριωμένου υποκατάστατου ήταν 87,2 ± 2,1% για το Ph-d10 και 90,3 ± 1,8% για το Chy-d12, αντίστοιχα (n = 3). Αιχμηρά δείγματα σε κάθε παρτίδα αναλύθηκαν με μέσο όρο ανάκτησης 86,7 ± 2,6% για το Ph και 89,6 ± 1,4% για το Py, αντίστοιχα (n = 3). Κάθε εκχύλισμα αναλύθηκε σε διπλή μορφή και οι σχετικές τυπικές αποκλίσεις ήταν μικρότερες από 20%. Τυχόν αναλύσεις που δεν πληρούσαν τις απαιτήσεις διασφάλισης ποιότητας αναλύθηκαν εκ νέου. 2.5. Στατιστικές αναλύσεις Όλα αυτά τα πειράματα πραγματοποιήθηκαν εις τριπλούν και τα αποτελέσματα που παρουσιάστηκαν ήταν μέσες τιμές των τριών επαναλήψεων. Τα δεδομένα αναλύθηκαν στατιστικά χρησιμοποιώντας ανάλυση διακύμανσης (ANOVA) και χρησιμοποιήθηκαν οι δοκιμές πολλαπλών εύρους του Duncan για να προσδιοριστεί η σημασία των διαφορών μεταξύ των παραμέτρων. ο

265

Table 1. Phenanthrene and pyrene removal rates in different sampling zones in planted and unplanted treatments after 60 days of K. candeli growth. Zones

Treatment Phenanthrene (%) Unplanted

S0 S1 S2 S3 S4

27,1 26,5 25,6 26,1 25,5

± ± ± ± ± ±

3,5Cb 3,4Cb 3,1Cb 2,9Cb 3,2Cb

Grain (%) planted 47.5 53.6 56.8 43.2 37.3

± ± ± ± ± ±

Uplantet 4.1Aa 6.4Aa 6.2Aa 3.4Aa 3.2Ba

23,5 24,4 21,9 23,2 20,9

± ± ± ± ± ±

2.9Cb 2.6Cb 2.3Cb 2.8Cb 2.6Ca

Planted 32.4 46.2 47.7 38.1 23.5

± ± ± ± ± ±

3,7Ba 5,1Aa 4,9Aa 4,2Ba 2,6BCa

Note: Values in each column followed by different capital letters (A, B, C and D) indicate significant differences (p < 0.05) between different distances (0, 1, 2, 4 and 6 mm) from the root and in each series followed by different lowercase letters (a and b) indicates a significant difference between planted and unplanted soil using Duncan's multiple statistical soil tests. Values represent mean ± standard deviation. S0–S4 represents the distance of 0, 1, 2, 4 and 6 mm from the root surface.

The statistical package used was the SPSS statistical software package (version 11.0) and the confidence limit was 95%. 3. Results and discussion 3.1. Diffusion gradient of Ph and Py in the sediment At the beginning of the experiment, 10 ± 0.5 and 10 ± 0.4 mg kg−1 added Ph and Py in the sediment slurry were adsorbed on the sediments. This indicates that acetone evaporation did not cause significant loss of pungent PAHs in the sediment slurry. At the end of the 60-day experiment, the results showed that the initial concentrations of Ph (10.0 mg kg-1) and Py (10.0 mg kg-1) were significantly reduced in the planted sediment, as well as in the unplanted control, but a more significant extinction rate it was obvious when the plants were introduced. The removal rates of Ph and Py were 37.3–56.8% and 23.5–47.7%, respectively, in different zones of the sediment gradient planted with sediment, which were significantly higher compared to the non-planted treatments (25.5–27 .1% for Ph and 20.9-24.4% for Py ) (Table 1). The PAH concentration in the sediment after mangrove growth was influenced by the proximity of the roots. In both Ph- and Py-enriched sediments with plant treatments, the general trend of PAH degradation was typically rhizosphere > compartment > remote rhizosphere, except for a small difference in the root zones (Table 1). The mass balance results indicated that although the added PAHs were firmly adsorbed to the sediments at the beginning of the experiment, the PAH digesters had the ability to efficiently utilize and degrade the adsorbed PAHs (Table 2). Loss of PAHs from mangrove sediment may be a consequence of biotransformation, Table 2. Mass balance of phenanthrene and pyrene in rhizosphere sediment after 60 days of K. candeli growth. PAH (mg)

There is no vegetation

Fitoprerada

Intake of leachate Absorption in facilities Retention in sediment Loss

100,0 ± 5,0 ND NA 74,2 ± 4,5 25,8 ± 1,7

100.0 ± 5.0 ND NA 59.8 ± 3.7 40.2 ± 2.3

Intake of leachate Absorption in facilities Retention in sediment Loss

100,0 ± 4,0 ND NA 78,4 ± 5,2 21,6 ± 1,2

100,0 ± 4,0 ND NA 72,3 ± 4,4 27,7 ± 1,8

Note: ND, not found. NA, not relevant because no plants grew in the rhizosphere or PAH uptake was negligible. Values represent mean ± standard deviation.

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Planted phenanthrene

Cmic (mg kg-1)

biodegradation, plant uptake or abiotic diffusion, including leaching and volatilization [21,22]. In this study, abiotic leaching losses were negligible because no leachate was generated during the experiment. Loss of Ph and Py by volatilization from sediments is also unlikely due to total water coverage and low vapor pressures of PAHs (10−1.00 and 10−2.05 L atm mmol−1 for Ph and Py). Our data showed that the mangrove plant accumulates only a few PAHs (detailed in Section 3.4), therefore the loss of PAHs from the soil due to plant uptake/accumulation can be considered negligible. In the rhizosphere, Reilley's results suggest that abiotic diffusion (chemical decomposition and irreversible sorption) is not a possible loss pathway for anthracene and Py [23]. Therefore, our results indicated that increased diffusion of PAHs may be due to increased rhizosphere microbial density and activity compared to unplanted soil, as root exudates and plant debris could increase pollutant bioavailability, provide more substrates for co-metabolic degradation, and modify soil environment to make it more suitable for microbial transformation. Roots are known to release certain organic compounds, such as amino acids, organic acids, sugars, enzymes and complex carbohydrates, providing a source of carbon and energy for the growth of rhizosphere microorganisms [13,24]. Increased diffusion of PAHs in the rhizosphere may also be a consequence of reduced ability to extract PAHs by forming bound residues. The rhizosphere could stabilize pollutants by polymerization reactions such as hydration [12,25]. PAH degradation gradients observed in the root box showed that the diffusion of Ph and Py was higher in the sediment that received only root exudates from soil with root exudates and plant roots. This is inconsistent with the gradients in root exudates and plant enzymes or the most diffusion-limited zones of metal nutrient depletion in many reports [26,27]. This is an important but interesting conclusion. This may be the result of competition between plant roots and soil microbes for soil nutrients that affects the activities of soil microbes, especially in soils with low organic matter, such as the sediment used in this study. In addition, Ph subhydration caused by root senescence can make PAHs more hydrophobic and potentially affect their availability via adsorption [ 28 ]. Although the PAHs accumulated in the plants represented only a small amount of PAHs removed, it is not known whether this K. candel plant could produce enzymes to degrade PAHs. However, the synergistic mechanism has yet to be confirmed by further studies, which should include more plants and study different plant species.

Vegetative

12 8 4 0 S0

Nuclear

Cmic (mg kg-1)

266

Sown

Vegetative

12 8 4 0 S0

Fig. 2. Amount of carbon from microbial biomass (Cmic) at different distances near the roots of K. voska growing in sediment treated with phenanthrene and pyrene. Bars are standard error of the mean of three replicates.

of sediments with plants compared to unplanted treatments. In PAH-treated sediment, enzyme activity was highest in the rhizosphere or root compartment and then decreased with increasing distance from the root surface (Figures 2-5). However, enzyme activities did not decrease with distance in unplanted soils. This is in good agreement with the PAH degradation data (Table 1). The most likely number of PAH degraders was influenced by planting regime. Our data on microbial biomass support the hypothesis that microorganisms are responsible for the observed degradation of PAHs. Planted delicacy-

3.2. Sediment Microbial Biomass and Enzyme Activities Carbon content in soil microbial biomass and soil dehydrogenase, urease and phosphatase activities were measured to assess the effect of the rhizosphere gradient on PAH degradation. Sediments with different slopes from the roots showed different responses to the presence of PAHs in the root field. Overall, in the unplanted sediment the microbial biomass measured as total Cmic was the same in the different compartments but was lower than the planted sediment (Figure 2). Similarly, Cmic was 16-234% higher with plants than without. The highest concentration of Cmic (12.65 mg kg-1 for Ph and 10.68 mg kg-1 for Py, respectively) in each gradient zone was found in the rhizosphere (S1). Soil dehydrogenase, urease, and phosphatase activities in planted soils were higher than those in non-planted treatments during the 60-day biodegradation process (Figures 3-5). In our research, it was shown that relatively lower concentrations of Ph and Py (10.0 mg kg−1) have a stimulating effect on enzyme activity in the sediment. However, other researchers have found that higher concentrations of PAHs can inhibit the enzyme in soil [10]. Our data showed that rhizosphere effects caused increased response characteristics to

Fig. 3. Urease activities at different distances near the roots of K. voska growing in sediment treated with phenanthrene and pyrene. Bars are standard error of the mean of three replicates.

Η. Prema Lu i sur. / Journal of Hazardous Materials 196 (2011) 263–269

267

Table 3. Linear regression between residual PAH concentrations (Y) in the sediment and various biological parameters in the rhizosphere (S1) (Y = ax + b). Spiked PAHs

index (x)

fenantrene

Chemical urease dehydrogenase phosphatase

−0,36 −0,15 −5,76 −0,03

8,79 13,43 10,01 8,50

0,792* 0,693* 0,749* 0,508*

Nuclear

Chemical urease dehydrogenase phosphatase

−0,41 −0,17 −7,22 −0,04

9,37 12,69 10,05 9,77

0,645* 0,745* 0,761* 0,459*

* Significant (p < 0.05) difference between residual PAH concentration and biological parameters.

biological state of the soil [29]. In our experiment, although the urease and phosphatase activity responses were different and somewhat inconsistent with PAH degradation, it could be concluded that there was increased urease and phosphatase activity in planted soil, especially in the rhizosphere, compared to bare soil. The reason for different enzyme activities in different soil zones may be related to the gradient effect of root exudation. Fig. 4. Dehydrogenase activities at different distances near roots of K. voska growing in sediment treated with phenanthrene and pyrene. Bars are standard error of the mean of three replicates.

3.3. Correlation between microbial activities and PAH diffusion in the rhizosphere

The sites, especially the rhizosphere, contained significantly increased and large microbial biomass that could mediate increased PAH degradation. The observed differences between soils with and without plants, as well as between different sampling zones near the roots of the planted soil, were expected based on microbial growth and community structure modified by PAHs and root exudates. Total microbial activity determined by dehydrogenase, urease and phosphatase activities is an indicator

The successful application of root remediation is highly dependent on the ability of contaminant-degrading or plant growth-promoting microbes to effectively colonize growing roots. Table 3 shows the relationship between microbial activities and PAH diffusion in the rhizosphere after 60 days of cultivation. A significant negative correlation was found between the concentrations of residual pollutants and soil enzymes in the rhizosphere. Statistical correlations (r2) of both acute PAHs, especially for the two indices Cmic and dehydrogenase, had better values (rC2 = 0.792 and mic)

2 0.645, p < 0.001; rdehydrogenase = 0.749 and 0.761, p < 0.001). For phosphatase, it showed a relatively worse correlation. Some plant species appear to increase the number of degrading microbes in a large volume of soil that extends beyond the rhizosphere. The release of compounds or enzymes from the roots is thought to be related to rhizosphere biodegradation, and plant species differ in the nature and amount of released compounds, so the plant species used can be an important factor affecting the effectiveness of phytoremediation. Parrish et al. [ 30 ] reported that after 12 months of plant growth, PAH-degrading microbial populations in vegetated treatments were more than 100-fold greater than those in unvegetated controls. This microbial consortium can provide plants with various benefits, including the synthesis of compounds that protect plants by reducing plant stress hormone levels. Chelating agents to provide essential nutrients for plants. protection against plant pathogens. and degrading pollutants before they can adversely affect plants [31]. Therefore, the differences between rhizosphere soil and non-rhizosphere soil could be explained by the rhizosphere effect.

3.4. Ph and Py accumulation potential in plant tissue

Fig. 5. Phosphatase activities at different distances near roots of K. voska growing in sediment treated with phenanthrene and pyrene. Bars are standard error of the mean of three replicates.

When sharp sediments are used in remediation trials, the focus has often been on the ability of a given plant to accumulate a particular compound and can be removed with the biomass for sequestration or burning. To gain a comprehensive understanding of PAH degradation mechanisms,

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Table 4 Concentrations of phenanthrene and pyrene (mg kg−1) and concentration factors (CF) in the plant after 60 days of plant growth. Treatments

stick

RCF

stem

SCF

Ply

LCF

Phenanthren Pyren

0,83 ± 0,11 1,56 ± 0,25

0,18 0,29

0,55 ± 0,07 1,83 ± 0,24

0,12 0,34

0,32 ± 0,04 0,59 ± 0,07

0,07 0,11

The intake of Ph and Py by K. candel was measured (Table 4). Ph and Py concentrations in roots were higher than in stem and leaves, and Ph concentrations in leaves were lower among plant tissues (Table 4). Plant concentration factors (CF) were calculated as the ratio of PAH concentrations in plant tissue (root, shoot and leaf) and sediment based on dry weight. The results also showed that the root concentration factors (RCF) for Ph (0.18) were much lower than for the Py treatment (0.29). This can be explained by the higher Kow value (octanol-water partition coefficient) of Py than Ph [32]. It has been shown that hydrophobic compounds with log Kow > 4 are not easily absorbed by plants through transpiration due to their hydrophobicity. Log Kow for Ph and Py was 4.17 and 5.13, respectively [33]. Our data showed that K. candel is not a hyperaccumulating entity for PAHs (CF from 0.07 to 0.34). There were no significant correlations between the concentration of Ph and Py in the roots and the diffusion of Ph and Py from rhizosphere sediments (S1) as well as other gradient sediments (S2-S4). This indicated that the accumulation of Ph and Py in the roots is not the main factor contributing to the removal of PAHs from the soil. This result is similar to a previous report [ 34 ], which showed that the contribution of plant accumulation and uptake to the removal of Py from contaminated sediments is negligible. 4. Conclusion We investigated the degradation gradient of the mangrove K. candel rhizome for PAH-contaminated sediment. The presence of the mangrove plant significantly increased the diffusion of Ph and Py into the contaminated sediment. In the process of removing Ph and Py, there was a significant effect of the proximity of the roots, which depended on the distance from the root surface. The increased rate of dispersion in different gradients of planted compared to unplanted sediment was 11.8–29.9% for Ph and 2.9–25.8% for Py. The accumulation of PAHs in plant parts showed a negligible contribution to the total return. Diffusion promotion by plant roots was the dominant contributor to enhanced sediment remediation of Ph and Py in the presence of K. candel. Our results show that the increase in the disappearance of Ph and Py is caused by an increase in the biological activity of the rhizosphere compared to the sediment without roots. In addition, there is scope for future work, particularly regarding the underlying mechanisms responsible for the observed root regeneration effects. These future directions include elucidating the complex processes at the soil-microorganism-root interface. Greater efforts should be made to investigate the root exudates deposited in the rhizosphere and the microbial activities involved during the restoration process, and the results that can be achieved using mangroves in field trials. Acknowledgments This study was supported by the China Postdoctoral Science Foundation Project and the National Natural Science Foundation of China (10805036, 20777062, and 30710103908). The authors would like to thank Bosen Weng and Yong Huang for their assistance in sampling work and Dr. Youwei Hong for PAH analysis;

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Journal of Hazardous Materials 196 (2011) 270-277

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Amphiphilic hollow carbon microspheres for the sorption of phenol from water Zhengrong Guan, Li Liu, Lilu He, Sen Yang ∗ School of Resources and Environmental Sciences, Biomass Engineering Center, China Agricultural University, Beijing 100193, People's Republic of China

article

i n f o

Article history: Received June 29, 2011 Received in revised form August 19, 2011 Accepted September 7, 2011 Available online September 14, 2011 Keywords: Carbon spheres Phenol separation removal

Amphiphilic porous hollow carbon spheres (PHCS) were synthesized by mild hydrothermal treatment of yeast cells and further pyrolysis after treatment. The morphology, chemical composition, porosity and structure of carbonaceous materials were investigated. Obviously, the carbonaceous materials consisted of carbonaceous organic matter (COM) and non-carbonized organic matter (NOM), and the relative proportions of COM and NOM could be adjusted by changing the temperature of the hydrothermal and/or pyrolysis treatment. Phenol sorption properties of carbon materials have been investigated and the adsorption isotherms correspond well to the modified Freundlich equation. The sorption isotherm of phenol on PHCS was found to be practically linear even at extremely high concentrations, which is less reported for activated carbon or other inorganic materials. This type of adsorption isotherms is attributed to the distribution mechanism, and the highest value of the distribution coefficient (Kf ) and carbon-normalized Kf (Koc ) is 56.7 and 91.5 mL g−1, respectively. In addition, PHCS show fast adsorption kinetics and easy regeneration property. The results show that PHCS are potentially effective sorbents for the removal of unwanted organic chemicals in wastewater, especially in high concentrations. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Phenol-like compounds have caused worldwide concern due to their toxicity and the frequency and amount of their presence in wastewater from various industries, such as refineries (6–500 mg L−1), coking (28–3900 mg L ). -1), coal processing (9-6800 mg L-1) and petrochemical production (2.8-1220 mg L-1) [1-4]. Phenolic wastewater usually contains several different pollutants, and the concentration varies from a low concentration of a few mg L−1 to a high concentration of several thousand mg L−1. These concepts, the best wastewater phenol abatement technologies to be used, depend to a large extent on individual cases, especially the concentration of phenol in the stream, the presence of other pollutants, the nature of the plant where this problem exists. [2. ]. Now, numerous strategies such as ozone/hydrogen peroxide oxidation [5], biological methods [6,7], membrane filtration [8], ion exchange [9], electrochemical oxidation [10], photocatalytic degradation [11] and adsorption [ 12] were used to remove phenol. A review of available technologies for the removal of phenol from liquid streams was recently published, which compares the experimental conditions and performance of different techniques [2]. Adsorption is generally considered a

∗ Corresponding author. Phone: +86 10 62733470; fax: +86 10 62733470. E-mail address:[email protected](S. Mladi). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.025

an operationally simple, effective and widely used procedure for the removal and recovery of phenol. Among a large number of different adsorbents, activated carbon (AC) is the most commonly used adsorbent in industrial scale and experimental research [1]. However, AC is not an ideal adsorbent for practical applications at high phenol concentrations, since the adsorption amount of AC will soon reach saturation. Demanding regeneration and weak mechanical stiffness of AC are also problems for its wider application [13]. Consequently, a wide variety of non-conventional adsorbents have been investigated for the removal of phenols and phenolic pollutants, including organols [14], polymer [13], carbon nanotubes [15], sludge-based adsorbents [16] and activated carbon [16 ] 17] . However, the development of new adsorption materials for the removal and recovery of phenols and highly concentrated phenolic pollutants remains a major challenge. Yeast, a by-product of the brewing industry, is considered an industrial organic waste of great concern [18]. In a previous study, we reported a simple method for the fabrication of porous hollow carbon spheres (PHCS) with controlled shell porosity from Saccharomyces cerevisiae (S. cerevisiae) cells [19,20]. By mild hydrothermal treatment of these microscopic unicellular organisms, hollow microspheres with controlled meso- and macroporous shells were synthesized. The most interesting thing is that it was discovered that the surfaces of these hollow spheres are covered with both hydrophobic and hydrophilic functional groups, giving the obtained microspheres an amphiphilic property. In fact, we found that PHCS can be well distributed not only in water, but also in water

Z. Guan i sur. / Journal of Hazardous Materials 196 (2011) 270–277

σε μη πολικούς διαλύτες όπως το τολουόλιο και το χλωροφόρμιο. Αυτό μας ενέπνευσε να υποθέσουμε ότι τα PHCS μπορεί να είναι πολλά υποσχόμενο ροφητικό για οργανικούς ρύπους από υδατικό διάλυμα. Εδώ, τέσσερα είδη PHCS συντέθηκαν μέσω ήπιας υδροθερμικής επεξεργασίας κυττάρων ζυμομύκητα και περαιτέρω πυρόλυσης μετά την επεξεργασία. Και διερευνήθηκαν η μορφολογία, η χημική σύνθεση, το πορώδες και η δομή των ανθρακούχων υλικών. Εξετάστηκε η συμπεριφορά ρόφησης (δηλαδή, ισόθερμες προσρόφησης, μοντέλα κινητικής και η επίδραση του pH και της θερμοκρασίας) των PHCS προς τη φαινόλη. Αποσαφηνίστηκε η σχέση που υπάρχει μεταξύ της ειδικής επιφάνειας και της χημικής σύνθεσης των PHCS και της ικανότητάς τους να απορροφούν φαινόλη. Τέλος, διευκρινίστηκαν οι κύριοι μηχανισμοί προσρόφησης. 2. Πειραματικό 2.1. Σύνθεση PHCS Τα PHCS συντέθηκαν μέσω ήπιας υδροθερμικής επεξεργασίας κυττάρων ζυμομύκητα χρησιμοποιώντας τροποποιημένες μεθόδους που περιγράφηκαν στις προηγούμενες μελέτες μας [19]. Τυπικά, κύτταρα S. cerevisiae (3-4 g, αγορασμένα από την Angel Yeast Co., Ltd., Κίνα) προπλυμένα με ακετόνη διασκορπίστηκαν σε 2-3% (v/v) γλουταραλδεΰδη και αραιώθηκαν (λιγότερο από 0,01 mol L −1) υδατικό διάλυμα νιτρικού οξέος (40 mL), το οποίο στη συνέχεια τοποθετήθηκε σε αυτόκλειστο 50 mL σφραγισμένο με τεφλόν και διατηρήθηκε στους 180 ή 230 ◦ C για 8 ώρες. Τα στερεά προϊόντα του puce διαχωρίστηκαν φυγοκεντρικά, στη συνέχεια πλύθηκαν με τρεις κύκλους φυγοκέντρησης/πλύσης/επαναδιασποράς σε απιονισμένο νερό και αλκοόλη και ξηράνθηκαν σε φούρνο στους 80◦ C για 4 ώρες. Τα δείγματα PHCS που υποβλήθηκαν σε υδροθερμική επεξεργασία στους 180 και 230 ◦ C σημειώθηκαν ως P180 και P230, αντίστοιχα. Τα δείγματα που χαρακτηρίζονται ως P350 και P700 παρασκευάστηκαν μέσω περαιτέρω πυρόλυσης P180 σε θερμοκρασία 350 ◦ C και 700 ◦ C για 1 ώρα, αντίστοιχα, από έναν σωληνωτό αντιδραστήρα σε ροή αζώτου (15 mL min−1). 2.2. Χαρακτηρισμός PHCS Η μορφολογία των υλικών επιθεωρήθηκε με ηλεκτρονικό μικροσκόπιο σάρωσης πεδίου εκπομπής (SEM, JEOL, JSM-6700F, Ιαπωνία). Το εμβαδόν επιφάνειας και ο όγκος των πόρων των υλικών μετρήθηκαν με ισόθερμες προσρόφησης/εκρόφησης Ν2 στους 77 Κ με αναλυτή Physisorption (Micromeritics, ASAP 2020, U.S.A.). Η ανάλυση υπέρυθρων μετασχηματισμού Fourier (FTIR) πραγματοποιήθηκε με φασματόμετρο Micro FTIR (Nicolet, Magna 750 Nic-Plan FTIR Microscope, U.S.A.) στη φασματική περιοχή 4000–650 cm−1. Στερεάς κατάστασης διασταυρούμενης πόλωσης μαγική περιστροφή γωνίας και ολικής πλευρικής ζώνης καταστολής φάσματος πυρηνικού μαγνητικού συντονισμού άνθρακα 13 (13 C NMR) (CPMAS-TOSS) ελήφθησαν με φασματόμετρο Bruker Avance 400 MHz (Καρλσρούη, Γερμανία). Οι περιεκτικότητες C, H, N, O των δειγμάτων προσδιορίστηκαν χρησιμοποιώντας στοιχειακό αναλυτή (Elementar, Vario EL, Γερμανία). Η ατομική αναλογία (O + N)/C, O/C, H/C υπολογίστηκε με την περιεκτικότητα του στοιχείου. 2.3. Πειράματα ρόφησης Τα πειράματα παρτίδας διεξήχθησαν χρησιμοποιώντας μια σειρά φυγοκεντρικού σωλήνα με βιδωτό καπάκι 15 mL καλυμμένου με φύλλα τεφλόν για να αποτραπεί η εισαγωγή οποιασδήποτε μόλυνσης από ξένα σωματίδια. Σε τυπικό πείραμα παρτίδας, 20 mg του ροφητικού προστέθηκαν σε 10 mL διαλύματος φαινόλης σε διάφορες συγκεντρώσεις (0–10.000 mg L−1) που ελήφθησαν σε σφραγισμένους σωλήνες, οι οποίοι τοποθετήθηκαν στο συγκρότημα ανακίνησης του θερμοστάτη. Τα διαλύματα ανακινήθηκαν στις 150 rpm και σταθερές θερμοκρασίες για 24 ώρες για να επιτευχθεί εξισορρόπηση. Μετά την ισορροπία, τα μείγματα διηθήθηκαν μέσω φίλτρων σύριγγας νιτροκυτταρίνης 0,45 ␮m. Οι συγκεντρώσεις φαινόλης στα διηθήματα προσδιορίστηκαν με φασματόμετρο UV (Persee, TU-1810, China) στο μέγιστο μήκος κύματος προσρόφησης φαινόλης (270 nm) και pH 6 (pH ρυθμισμένο με 0,5 M

271

HCl or NaOH). The isotherms were derived by taking different concentrations of phenol at the calculated temperatures and pH values. All experiments were performed in triplicate and average data are reported. Standard deviations were found to be within 2.0%. In addition, the error bars for the figures were smaller than the symbols used to plot the graphs and are therefore not shown in the figures. 2.4. Sorption models and statistical analysis The Freundlich model was used to fit the sorption data in this work, mainly because of its advantages in investigating isothermal nonlinearity. In order to facilitate direct comparisons of sorption affinities between tested samples and to investigate the effect of temperature on sorption, a modified Freundlich equation was used to fit the sorption data in this paper: log qe = log KF + n log Cr Cr =

Ce Sw

(1) (2)

where qe is the concentration of the solid phase (mg g-1), Ce is the equilibrium concentration of the liquid phase (mg L-1). Sw is the solubility of phenol for a given temperature (mg L−1), and Cr is dimensionless since the value of Sw is constant for a given temperature and is expressed in the same unit as Ce. K F and n are modified Freundlich adsorption parameters, K F (mg g−1 ) is the adsorption capacity coefficient, which represents the mass of adsorbed phenol per unit mass of the sorbent when the Ce concentration approaches saturation, and n (without dissociation) is an indicator of isothermal nonlinearity associated with the heterogeneity of adsorption sites [21]. The distribution-adsorption model was also analyzed to describe adsorption from aqueous solutions to heterogeneous solids [14,22]: QT = QA + QP

(3)

where QT is the total amount of phenol adsorbed on the sorbent. QA and QP are the amounts contributed by absorption and distribution. According to the distributive adsorption model, the distributive effect is gradually favored by increasing solute concentration, while the adsorption contribution reaches saturation faster with solute concentration. The isotherm at high concentrations should approach linearity. Therefore, adsorption in the region of high solute concentration becomes saturated and remains a linear distribution. So, equation (3) can be converted into: QT = QAmax + QP = QAmax + Kf Ce Koc =

Brand Kf

(4) (5)

where QAmax is the saturated adsorption capacity estimated from high concentration data. Kf Ce is the distribution contribution at high concentration, where Kf is the distribution coefficient. Ce is the equilibrium concentration of the dissolved substance (mg L−1). Koc is carbon normalized to Kf, and foc is the percentage of carbon content in the sorbent. Linear regression between QT and Ce was performed in the high solute concentration range, and QAmax corresponds to the y-intercept of the line and Kf to the slope. 2.5. Kinetics of sorption Batch technique was used for kinetic studies. Typically, 20 mg of PHCS was added to 10 mL of phenol solution (2000 mg L-1) in a series of screw-capped centrifuge tubes and then shaken on a constant temperature rotary shaker (150 rpm) at 25 ± 0.5 ◦ C.

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sl. 1. SEM slike (a) P180, (b) P230, (c) P350 i (d) P700.

The concentrations of phenol in the solution were sampled and analyzed at different time intervals, and the data were averaged. Pseudo-first and pseudo-second order linear models are given by the following equations [23,24]: ln(qe − qt ) = ln qe − k1 t

(6)

t 1 t = + qt qe k2 q2e

(7)

t1/2 =

1 x2 this

(8)

h = k2 q2e

(9) mg g−1 )

where qe and qt (both amounts of adsorbed phenol per unit mass of sorbent at equilibrium and time t (h), respectively); k1 (h−1) and k2 (g mg−1 h−1) are pseudo-, first-order and pseudo-second-order rate constants, respectively. For the pseudo-second-order kinetic model, the sorption half-time (t1/2) and the initial sorption rate (h) are given (equation (8) and (9)). The value t1/2 is the time required to take up half of the maximum absorbed amount of sorbate in equilibrium 3. Results and discussion 3.1 The characteristics of PHCS SEM images and optical microscopic photographs of carbonized products are shown in Figures 1 and S1. The hydrothermal products (P180 and P230) were found to be porous hollow microspheres in the size range of 2.0–4.0 m, consistent with

with our previous results [19]. The thermal stability of the microspheres was satisfactory. After post-crack treatment of P180 at 350 ◦ C, the structure of the microspheres is still well preserved (Fig. 1c for P350). However, a further increase in the pyrolysis treatment temperature to 700 ◦ C changed the morphology of the material from microspheres to larger carbon blocks (Fig. 1d for P700). Surface area and pore volume of the material were measured by N2 adsorption/desorption isotherms at 77 K, and all BET surface areas of the material are below 10 m2 g−1 (table 1). The nitrogen adsorption/desorption isotherm of the material at 77 K with the corresponding pore size distribution is shown in the figure. S2. Fig. Fig. 2 shows the solid state 13C NMR spectra of the carbon products. The peaks at ı = 26 and 31 ppm can be attributed to methyl and methylene, respectively, and those in the range ı = 120-150 ppm can be attributed to long-range conjugated C-C bonds and oxygen-substituted C-C bonds, revealing the existence of a composite aromatic furan ring [25, 26]. . In addition, oxygenated functional groups, including carbonyl, carboxy, hydroxy, ether and ester groups, are also disclosed. The content of aromatic species was significantly improved by pyrolysis treatment of hydrothermal products, which indicates a high degree of carbonization for P350, and especially for P700. In addition, the average chemical shift decreases with the carbonization sequence, which is similar to the coordination pattern of graphite. The corresponding Fourier transform infrared (FTIR) spectra of the carbon materials are shown in the figure. 3. Both aliphatic and aromatic species were detected for P180, P230 and P350. The bands at 2925, 2861, 1442 and 1372 cm−1 are mainly attributed to CH2 units [27]. Those at 1693 and 1160 cm−1 are attributed to CO and C–O stretching vibrations of ester bonds. And the band at 1602 cm−1 is assigned to C C and C O

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Table 1 Elemental compositions, atomic ratios, BET surface area (SA) and total pore volume (TPV) of carbon materials. Sample

Processing temperature (◦ C)

C (% by weight)

H (wt%)

O (wt%)

N (% by weight)

Ratio (O + N)/C

O/C system

H/C system

SA (m2 g−1 )

TPV (ml g-1)

P180 P230 P350 P700

180 230 350 700

62,00 72,96 73,22 77,17

6,25 6,99 4,44 1,94

21,16 12,15 10,61 5,71

6,08 4,88 6,27 5,06

0,34 0,18 0,18 0,11

0,26 0,13 0,11 0,06

1,21 1,14 0,72 0,30

9,504 2,635 1,960 2,830

0,0498 0,0095 0,0052 0,0080

H/C: atomic ratio of hydrogen to carbon. O/C: atomic ratio of oxygen to carbon. (O + N)/C: atomic ratio of the sum of nitrogen and oxygen to carbon.

Fig. 2. Solid state 13C NMR spectrum of (a) P180, (b) P230, (c) P350 and (d) P700.

stretching in the aromatic ring [22]. The peak at 786, 706 cm−1 can be attributed to out-of-plane aromatic CH distortion. In the case of P700, almost all of the above band intensities are dramatically reduced or disappear, indicating destruction

surface functional groups and chemical structure after pyrolysis treatment at high temperatures. The elemental compositions shown in Table 1 are in good agreement with FTIR observations. The content of O decreased from 21.16% for

sl. 3. FTIR spektar (a) P180, (b) P230, (c) P350 i (d) P700.

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Fig. 4. Phenol sorption isotherms on carbonates in aqueous solution at 25 ◦ C and pH 6, inset: P230 sorption isotherms at high phenol concentrations.

P180 to 12.15% for P230 and further decreased slightly to 10.61% for P350. When the pyrolysis treatment temperature increased to 700 ◦ C, the O content drastically decreased to 5.71% (P700). The H content in P350 and P700 decreased dramatically from about 6.25% (P180) to 4.44% and 1.94%, respectively. These results mean that the degree of carbonization of the samples increased with higher hydrothermal temperature and pyrolysis temperature. The H/C ratio and O/C ratio decrease with deep carbonation, indicating that product surfaces become less hydrophilic [28]. The decrease of the polarity index [(O + N)/C] with the degree of carbonization reveals the decrease of the polar functional groups of the surface [22]. All the above results show that both hydrothermal and pyrolysis treatments have changed the chemical composition of PHCS to a large extent. Due to the partial loss of hydrophilic groups and aromatization of the molecular frameworks, it is assumed that the wetting properties of PHCS are very different from the original hydrophilic yeast cells. We found that all samples except P700 could be well dispersed not only in water, but also in nonpolar solvents (fig. S3), suggesting that PHCSs have amphiphilic surfaces, as we previously reported [ 19 , 20 ]. In addition, there are two types of carbon in PHCS, namely alkyl carbon and aromatic carbon. Alkyl carbon originates mainly from non-carbonized organic matter (NOM) of yeast cells, and aromatic carbon is a carbonization product, i.e. carbonized organic matter (COM). It is obvious that the relative proportions of COM and NOM can be adjusted by changing the temperature of the hydrothermal and/or pyrolysis treatment. Although there are large amounts of NOM in samples P180, P230 and P350, COM is dominant in P700. The amphiphilic nature and special chemical composition of PHCS make them excellent potential sorbents for the removal of organic compounds from wastewater. Then, phenol adsorption isotherms were derived. 3.2. Sorption isotherms The sorption of phenol on carbon materials at 25 ◦ C was tested, and the results are shown in Figure 4. The sorption of P700 was saturated at a very low concentration of phenol (86 mg L−1), and the amount of sorption was only 5.5 mg g-1, which may be a consequence of the very small surface area (2.8 m2 g-1). Since P700 was produced by pyrolysis of P180 at 700 ◦ C for 1 hour, the high degree of carbonization and several surface functional groups (Figures 2 and 3) suggest that the sorption mechanism of P700 is physisorption.

Το αποτέλεσμα σημαίνει ότι το P700 δεν μπορεί να χρησιμοποιηθεί ως ροφητικό για οργανικές ενώσεις, και έτσι δεν συζητήσαμε πια για το P700. Το εμβαδόν επιφάνειας των P180, P230 και P350 ήταν όλα μικρότερα από 10 m2 g−1, ωστόσο, εμφάνισαν αρκετά διαφορετικές συμπεριφορές προσρόφησης: ποσότητα ρόφησης ανεξάρτητη από το εμβαδόν επιφάνειας και σχεδόν γραμμική ισόθερμη χωρίς κορεσμένη προσρόφηση. Αυτό το χαρακτηριστικό προσρόφησης αναφέρθηκε λιγότερο για το AC [29,30] ή άλλα ανόργανα προσροφητικά υλικά [31], γεγονός που υποδηλώνει ότι ο μηχανισμός προσρόφησης των PHCS δεν είναι απλή φυσική ή χημική προσρόφηση. Το τροποποιημένο μοντέλο Freundlich εφαρμόστηκε για τη διερεύνηση της ισόθερμης μη γραμμικότητας των PHCS. Οι ισόθερμες ρόφησης ταιριάζουν καλά με την τροποποιημένη εξίσωση Freundlich και οι υπολογισμένες παράμετροι παρατίθενται στον Πίνακα 2. Η ισόθερμη της φαινόλης στο P180 είναι πρακτικά γραμμική, με Freundlich n = 1,016 ± 0,025, και την εμφάνιση ισόθερμων P530 και διαφορετικής γραμμής για P5 κοίλη προς τα κάτω καμπυλότητα σε χαμηλές συγκεντρώσεις διαλυμένης ουσίας αλλά παρουσιάζουν πρακτικά γραμμικό σχήμα σε μέτριες έως υψηλές συγκεντρώσεις. Τα μη γραμμικά φαινόμενα είναι σχετικά πιο ορατά για το P350 παρά για το P230. Παρόμοια αποτελέσματα παρατηρήθηκαν για την απορρόφηση οργανικών διαλυμένων ουσιών από το νερό σε ένα ευρύ φάσμα Ce/Sw από μαύρους άνθρακες (BCs) ή βιοχαρακτήρες, που προέρχονται κυρίως από την ατελή καύση βιομάζας και ορυκτών καυσίμων [22,32,33]. Το μοναδικό ισόθερμο σχήμα, δηλ. μη γραμμικό σε χαμηλό Ce/Sw αλλά ουσιαστικά γραμμικό σε άλλα Ce/Sw, υποδηλώνει ότι περισσότεροι από ένας μηχανισμοί λειτουργούν σε όλο το εύρος συγκεντρώσεων [34]. Για να εξηγηθεί η μη γραμμικότητα των ισόθερμων προσρόφησης, προτάθηκαν μοντέλα προσρόφησης διπλού τρόπου (DMSM) ή μοντέλα διπλής αντιδραστικής περιοχής (DRDM). Σύμφωνα με αυτά τα μοντέλα [14,22,32,33,35], ο ροφητής θεωρήθηκε ως μια ετερογενής ουσία και μια έννοια NOM (ή «μαλακός άνθρακας», διογκωμένη, ελαστική κατάσταση) έναντι COM (ή «σκληρό άνθρακας», συμπυκνωμένη, υαλώδης κατάσταση) χρησιμοποιήθηκε για να οριοθετηθεί λειτουργικά η χημική ετερογένεια του ροφητικού και για να διευκρινιστούν οι μηχανισμοί για την προσρόφηση από τα εδάφη, τα ιζήματα, τους βιοενώσεις και τον άνθρακα. Το COM αναμένεται να συμπεριφέρεται ως φυσικός προσροφητής, παράγοντας ισόθερμη μη γραμμικότητα, και το NOM μπορεί να προσλάβει ρύπους μέσω ενός μηχανισμού διαχωρισμού (προσρόφησης) [22,32,33]. Ένα μοναδικό χαρακτηριστικό της διαδικασίας κατανομής είναι ότι η αναλογία των συγκεντρώσεων στερεάς φάσης προς υδατική φάση παραμένει αμετάβλητη με τη μεταβολή της συγκέντρωσης της διαλυμένης ουσίας. Με αυτή την έννοια, οι προσροφητικές προσλήψεις καθορίζονται από τα σχετικά ανθρακούχα και μη ανθρακούχα κλάσματα και τις επιφανειακές και χύδην ιδιότητες τους. Έτσι, η γραμμικότητα του P180 μπορεί να αποδοθεί στη ρόφηση της φαινόλης στο NOM μέσω μηχανισμού κατανομής, καθώς το NOM είναι το κυρίως είδος άνθρακα σε αυτό το δείγμα. Η μη γραμμικότητα των P230 και P350 σε χαμηλό Ce/Sw αποδίδεται σε μια συνδυασμένη φυσική προσρόφηση σε μια μικρή ποσότητα COM και σε μια επίδραση κατανομής της NOM. Σε μέτρια έως υψηλή Ce/Sw, η φυσική προσρόφηση γίνεται σε μεγάλο βαθμό κορεσμένη και η κατανομή στο NOM κυριαρχεί για να δώσει μια ουσιαστικά γραμμική ισόθερμη. Η μεγαλύτερη καμπυλότητα σε χαμηλό Ce/Sw του P350 οφείλεται πιθανώς στην υψηλότερη περιεκτικότητα σε αρωματικά τμήματα με αυξημένο βαθμό ενανθράκωσης. Η συμπεριφορά προσρόφησης του P700 με τον υψηλότερο βαθμό ενανθράκωσης, επικυρώνει αυτή την εικασία. Παρατηρήθηκε επίσης υψηλή ισόθερμη μη γραμμικότητα για χαρακτήρες υψηλής θερμοκρασίας [33]. Από το μοντέλο κατανομής-προσρόφησης, υπολογίστηκαν τα QAmax , Kf και Koc (Πίνακας 2), τα Kf και Koc αυξάνονται με την τάξη P350 < P230 < P180, η οποία είναι αντίστροφη με τον βαθμό ενανθράκωσης τους. Η φυσικοχημική φύση του οργανικού άνθρακα έχει προταθεί ως κύριος παράγοντας που ελέγχει την προσρόφηση οργανικών ενώσεων σε φυσικό ή τροποποιημένο οργανικό ροφητή. Σύμφωνα με δεδομένα FTIR και NMR (Εικ. 2 και 3), η κύρια φάση διαχωρισμού στα P180, P230 και P350 είναι ένα πολυμερές αλειφατικό κλάσμα που διατηρείται κατά τη διάρκεια της υδροθερμικής διεργασίας και η περιεκτικότητα του αλειφατικού κλάσματος αυξήθηκε της τάξης του P350< P230 < P180, το οποίο είναι σύμφωνο με το αυξημένο εφέ κατάτμησης. Τα παρόμοια συμπεράσματα παρουσιάστηκαν και από άλλους ερευνητές [22,36]. Οι Chefetz et al. [36] δοκίμασε τη ρόφηση του πυρενίου σε μια σειρά ροφητών που αποτελούνται από διαφορετικά επίπεδα αρωματικότητας και αλειφατικότητας. Σε αυτή τη μελέτη, α

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Table 2. Parameters of the modified Freundlich model, distribution coefficients and adsorption of saturated phenol on PHC. Sample

Modified parameter log of the Freundlich KF model a

P180 P230d P230 P230e P350

3,733 3,102 3,434 3,230 2,543

± ± ± ± ± ±

nb 0,049 0,024d 0,058 0,018e 0,041

1,016 0,661 0,822 0,700 0,378

± ± ± ± ± ±

0,025 0,013d 0,025 0,009e 0,019

R2 partition adsorption model

Kf (ml g−1)y

QAmax (mg g-1)γ

0,996 0,997d 0,987 0,991e 0,986

56.7 25.8d 38.2 39.4e 17.3

0 67,3d 45,8 34,3e 45,15

Koc (ml g-1)

maximum QA,SA (mg g−1 m−2)

91.5 35.4d 52.5 54.0e 23.6

0 25,5 17,4 13,0 23,0

max Koc is normalized to carbon Kf. QA,SA is the SA-normalized QAmax. The solubility of phenol in water at 15, 25 and 35 ◦ C is 8200 mg/100 mg, 8660 mg/100 mg and 9910 mg/100 mg, respectively. 95% confidence interval for log KF. b 95% confidence interval of n. c The slope and y-intercept of the linear equation were used to calculate the distribution coefficient (Kf ) and the maximum adsorption capacity (QAmax, mg g-1, Chen et al. [22]), respectively. d Tested at 15 ◦ C. e Tested at 35 ◦ C.

A positive trend was observed between the level of Koc and the aliphaticity of the pooled samples. Chen et al. [22] performed sorption experiments with biocompounds produced by pyrolysis of pine needles at different temperatures. This study clearly showed a higher adsorption affinity of naphthalene, nitrobenzene and m-dinitrobenzene on aliphatic-rich biospecies than on aromatic-rich biospecies. Table 2 shows the P230 and P350 maxima, which far exceed QA,SA by an amount corresponding to the small surface area of the sorbent. Some researchers have also reported higher adsorption of polar solutes compared to the small surface area of the sorbent. To explain the higher absorption of polar pesticides at low (relative) concentrations, Spurlock and Biggar [37] proposed a specific interaction model. The model assumes a specific interaction between polar solutes and highly active organic carbon phase sites. This means that the most active place, the most specific interaction and max. P350 showed lower oxygen content and very active higher QA,SA max due to dehydration by pyrolysis at 350 ◦ C, however, QA,SA for P350 is 23.0 mg g−1 m−2, higher than 17.4 mg g−1 m - 2 from P230. We hypothesized that the possible phenol-sorbent interaction may have a hydrophobic effect since P350 showed higher hydrophobicity compared to P230.

3.3. Kinetics of sorption Fig. Figure 5 shows the kinetics of phenol sorption on P180 and P230. Absorption rates were evidently very rapid. For example, given the test conditions, approx. 86 and 85% adsorption within 0.5 h for P180 and P230, respectively. In order to quantitatively compare the apparent adsorption kinetics between P180 and P230, the data were fitted to pseudo-first-order and pseudo-second-order models. The kinetic sorption constants are listed in Table 3. The regression coefficient (R2) of the pseudo-first-order model ranged from 0.9830 to 0.7022, and the Qe values calculated from the model differed greatly from the experimental values, which together indicate invalidity model. However, the pseudo-second-order model provided the best fit to all experimental data. The graphs show regression coefficients greater than 0.9997 for P180 and P230. The value of the constant k2 of P180 is greater than the value of P230, which is inversely proportional to Kf. Therefore, it can be assumed that the kinetics of phenol according to the pseudo-second-order model was governed by the adsorption process, and adsorption was the main rate-limiting adsorption step. 3.4. Effect of temperature on sorption In order to investigate the effect of temperature on phenol sorption, adsorption experiments were carried out at 15, 25 and 35 ◦ C for P230. The results are shown in fig. 6. Its parameters

Fig. 5. Sorption kinetics shown as adsorbed amount of phenol versus time at P180 and P230 with an initial phenol concentration of 2000 mg L−1 and 25 ◦ C.

Fig. 6. Phenol sorption isotherms on P230 at 15, 25 and 35 ◦ C at pH 6.

276

Z. Guan i sur. / Journal of Hazardous Materials 196 (2011) 270–277

Table 3. Adsorption rate constants for two kinetic models at 25 ◦ C and pH 6. Sample

P180 P230 a b

qe (exp)a (mg g-1)

98,63 107,85

First class model

Second class model

K1 (h−1)

qe (cal)b (mg g-1)

K2 (g mg−1 h−1)

qe (cal)b (mg g-1)

t1/2 (t)

h (mg g−1 h−1 )

3,17 0,12

82,92 19,86

0,9830 0,7022

0,1028 0,0660

110,31 106,80

0,099 0,14

1000 769

0,9997 0,9998

Experimental data. Calculated data from the model.

The modified Freundlich adsorption model and the distribution adsorption model are included in Table 2. The comparison of the modified Freundlich parameter n shows that the highest value is obtained at 25 ◦ C and the lowest value is obtained at 15 ◦ C. This means that the effective temperature of absorption and distribution are different. If we compare QAmax and Kf of P230 at 15, 25 and 35 ◦ C, we can find that the distribution increases with increasing temperature, which means that temperature can play an important role for the distribution coefficient of Kf. To date, a smaller number of Kf values have been reported for hydrophobic organic pollutants (HOCs) at temperatures other than 25 ◦ C, and the published conclusion is different [38]. Chen and Pawliszyn [39] evaluated the effect of temperature on Kf for BTEX compounds and found that Kf should not be strongly dependent on temperature. Muijs and Jonker [40] determined Kf values for several PAHs and found that temperature can play an important role. We hypothesized that the increasing Kf values of P230 for phenol with increasing temperature may be due to the polarity of phenol and the structure of P230. However, adsorption decreases with increasing temperature, indicating that the process is clearly exothermic.

3.5. Effect of pH on sorption The effect of pH on the adsorption capacity of P230 was tested, and the results are shown in the figure. 7. According to the model proposed by Deryło-Marczewska and Marczewski [41], the adsorption of a non-ionized compound does not depend on pH and surface charge. Adsorption isotherms at pH 3 and 6 coincided. This may be because the degree of dissociation of phenol is very low (pKa = 9.89) under acidic conditions. However, the sorption at pH 11 naturally decreased, because phenol is largely decomposed at pH 11 and dissolves in water.

3.6. Regeneration of sorbents Important goals in the development of sorbent materials include simple regeneration and isolation of sorbates [42]. Regeneration makes it possible to reuse absorbent material and reduce costs. Desorption of phenol from charged P230 using 0.01 M NaOH solution was found to be rapid. For example, after placing P230 adsorbed 141 mg g-1 phenol in 0.01 M NaOH solution for 10 min, the regenerated P230 showed that more than 98% of phenol was removed from the sorbent. Easy regeneration is due to the high solubility of the sodium salt of phenol in water. 4. Conclusions PHCS amphiphiles were synthesized by mild hydrothermal treatment of yeast cells and further pyrolysis after treatment. The sorption properties of PHCS for phenol in aqueous solutions were investigated, and the following conclusions were drawn: (1) PHCS are composed of COM and NOM, and the relative proportions of COM and NOM can be adjusted by changing the hydrothermal temperature and/or pyrolysis treatment . (2) The sorption isotherm of phenol on PHCS was practically linear even at extremely high concentrations. This type of adsorption isotherms is attributed to the distribution mechanism, and the highest value of the distribution coefficient (Kf ) and normalized carbon Kf (Koc ) are 56.7 and 91.5 mL g−1, respectively. (3) PHCS show fast adsorption kinetics and easy regeneration property. The results show that they are potentially effective sorbents for the removal and recovery of unwanted organic chemicals in water treatment, especially at high concentrations. Acknowledgments This work was supported by the National Natural Science Foundation of China (20703065, 20877097, and 20806089), the Ministry of Science and Technology of China (2008AA06Z324), and the Foundation for Scientific Universities of China (2011JS160). Appendix A. Supplementary data Supplementary data related to this article can be found in the online version at doi:10.1016/j.jhazmat.2011.09.025. bibliographical references

Fig. 7. Sorption isotherms of phenol on P230 at pH 3, 6 and 11 at 25 ◦ C.

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Distribution Behavior and Stabilization of Hydrophobically Coated HfO2, ZrO2, and Hfx Zr1−x O2 Nanoparticles with Natural Organic Matter Reveal Crystal Structure-Dependent Differences Divina A. Navarro 1, Sean W. Depner, David F. Wat Agson, Diana Sarbajit Banerjee ∗∗ Department of in Chemistry, University at Buffalo, State University of New York, Buffalo, NY 14260-3000, USA

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Article history: Received April 11, 2011 Received in revised form September 7, 2011 Accepted September 8, 2011 Available online September 14, 2011 Keywords: Double metal oxides Natural organic matter Phase transfer Nanoparticle structure Environmental mobility Colloidal interactions

a b s t r a c t Interactions between manufactured nanomaterials and natural organic matter (NOM) have a profound effect on the mobility of the former in the environment. The influence of specific structural features of nanomaterials on the distribution and colloidal stabilization of produced nanomaterials in different ecological spaces remains unexplored. Here, we present a systematic study of the interactions between humic acid (HA, as a model for NOM) with monodisperse, well-characterized, passive HfO2, ZrO2, and Hfx Zr1−x O2 in solid solution (NP). We note that mixing with HA causes almost complete phase transfer of hydrophobically coated monoclinic metal oxide (MO) NPs from hexane to water. Furthermore, HA appears to provide significant colloidal stabilization of the NPs in the aqueous phase. In contrast, phase transfer and colloidal stabilization of the aqueous phase were not observed for square MO-NPs. A mechanistic model for phase transport and aqueous dispersion of MO-NPs is proposed based on evidence from transmission electron microscopy, ␨-potential measurements, dynamic light scattering, Raman and infrared spectroscopy, elemental analysis and systematic experiments on tightly bound MO-NP ensembles. of different composition and crystal structure. The data indicate the synergistic role of the superlayer (micellar), ligand substitution (coordinative) and electrostatic processes in which HA acts both as an amphiphilic molecule and as a charged chelating ligand. The strong observed preference for phase transfer of monoclinic versus tetragonal NPs suggests the importance of preferential binding of HA to specific crystallographic faces and suggests the possibility of designing NPs to reduce their mobility in an aqueous environment. © 2011 Elsevier B.V. All rights reserved.

1. Introduction The upcoming large-scale commercialization of engineered nanomaterials (ENMs) has raised concerns about their possible environmental impacts [1-4]. Some preliminary data on the toxicity of ENMs at the subcellular, cellular, and organismal levels are becoming available [5-8]. Among these materials, transition metal oxide (MO)-based nanoparticles find various applications as nanoceramic fillers in composites, magnetic recording media, catalyst supports, and sensing elements [9]. More specifically, HfO2 and ZrO2 NPs have found wide application in optical and protective coatings

∗ Corresponding author. Phone: +1 716 645 4220; fax: +1 716 645 6963. ∗∗ Corresponding author. Phone: +1 716 645 4140; fax: +1 716 645 6963. E-mail addresses:[email protected](D.S. Aga),[email protected](S. Banerjee). 1 Present address: CSIRO Land and Water, Advanced Materials Transformational Capability Platform, Nanosafety, Biogeochemistry Program, Waite Campus, Waite Rd, SA 5155, Australia. 0304-3894/$ – see cover page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.028

technologies due to their thermal stability and high dielectric constant [10]. These materials are particularly promising alternatives to SiO2 as dielectric gate layers for flexible electronics [11-13]. The basic premise of flexible electronics is affordability and ubiquitous availability in standard media such as paper and fabric. As a result, the release of these NPs in various areas of the environment (especially in water and soil) is inevitable with the increase in commercial production, consumer use and end-user disposal. In particular, waste generated during production processes will have a high concentration of NP. Although the probability of release of MO NPs attached to device structures is not high, release into the environment can occur over a long period of time, especially as a result of material misuse and towards the end of product life. Given the low suggested price of flexible electronics and the lack of disposal specifications provided by manufacturers, many of these materials may eventually end up in landfills and runoff streams via household waste disposal. A systematic understanding of the distribution behavior, potential mobility and persistence of MO NPs is therefore crucial to

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αξιολόγηση πιθανών οικολογικών κινδύνων και διαμόρφωση τεκμηριωμένης πολιτικής [1]. Μέχρι τώρα, η περιβαλλοντική τύχη και η μεταφορά των ENMs έχουν χαρακτηριστεί κάτω από διαφορετικές περιβαλλοντικές συνθήκες (δηλαδή, pH, ιοντική ισχύς, οργανικά κολλοειδή κ.λπ.). Η επίδραση της φυσικής οργανικής ύλης (NOM) στη συμπεριφορά των NP στο περιβάλλον έχει τονιστεί σε πολλές μελέτες λόγω της πανταχού παρουσίας της πρώτης σε υδάτινα και εδαφικά περιβάλλοντα. Πράγματι, η φύση και η ποσότητα του NOM στο νερό έχει αποδειχθεί (τόσο θεωρητικά όσο και πειραματικά) ότι επηρεάζει τη σταθερότητα και τη βιοδιαθεσιμότητα μιας ποικιλίας NPs [14-20]. Σε πολλές περιπτώσεις, η σταθερότητα των NPs σε υδατικό εναιώρημα αποδίδεται συχνά στην προσρόφηση του NOM. Ωστόσο, αυτές οι αναφορές τείνουν να θεωρούν την προσρόφηση ως πρωταρχικό μηχανισμό αλληλεπίδρασης με βάση έμμεσα μέτρα όπως το ηλεκτρονικό μικροσκόπιο μετάδοσης (TEM), το ␨-δυναμικό και η σκέδαση φωτός, τα οποία είναι κάπως περιορισμένα στον χαρακτηρισμό αλληλεπιδράσεων που σχετίζονται με την επιφάνεια. Μια αξιολόγηση της χημικής δομής του NOM και των κολλοειδών MO-NPs προτείνει διάφορες διακριτές διεργασίες που μπορούν να διευκολύνουν τη σταθεροποίηση των NPs στην υδατική φάση. Το NOM έχει μια εξαιρετικά πολύπλοκη μοριακή δομή, η οποία περιλαμβάνει έναν σκελετό αλκυλικών και αρωματικών μονάδων με κρεμαστά λειτουργικές ομάδες που περιλαμβάνουν τμήματα καρβοξυλικού οξέος, φαινολικού υδροξυλίου και κινόνης [21-23]. Τα MO-NP, από την άλλη πλευρά, περιλαμβάνουν έναν ανόργανο πυρήνα που παθητικοποιείται από ένα στρώμα οργανικών προσδεμάτων [11,24]. Ο κρυσταλλικός πυρήνας μπορεί να υιοθετήσει διαφορετικές κρυσταλλικές δομές ανάλογα με τις ιδιαιτερότητες της σύνθεσης και της στοιχειομετρίας (τα NPs HfO2 και ZrO2 υιοθετούν μονοκλινικές και τετραγωνικές κρυσταλλικές δομές, αντίστοιχα) [25,11]. Οι εγγενείς μορφολογικές, ενεργητικές και επιφανειακές χημικές ιδιότητες του NOM τους επιτρέπουν να αλληλεπιδρούν και να σταθεροποιούν διαφορετικά είδη μέσω αμφιφιλικών διεργασιών και διεργασιών χηλίωσης μετάλλων. Εκτός από τους οργανικούς συνδέτες που περιβάλλουν τον κρυσταλλικό πυρήνα, τα MO-NPs έχουν επίσης εξαιρετικά αντιδραστικές επιφάνειες (άκρες και γωνιακές θέσεις) που επηρεάζουν σε μεγάλο βαθμό τη συμπεριφορά και την αντιδραστικότητά τους [26-28]. Ενώ αυτά τα δομικά και επιφανειακά χαρακτηριστικά είναι ευρέως γνωστό ότι είναι σημαντικά στην επιφανειακή επιστήμη, αυτές οι λεπτομέρειες συνήθως έχουν παραβλεφθεί σε πολλές μελέτες τύχης και μεταφοράς. Πολύ λίγες μελέτες μέχρι στιγμής έχουν επικεντρωθεί στους μετασχηματισμούς των παθητικοποιημένων από συνδέτη NP που παρασκευάζονται με μεθόδους θερμής κολλοειδούς χημείας κατά τις αλληλεπιδράσεις με το NOM. Τέτοια NP που καλύπτονται από πρόσδεμα είναι πράγματι πιθανό να είναι ο βασικός άξονας των περισσότερων τεχνολογιών που υποστηρίζουν τη νανοεπιστήμη [29-32]. Εδώ, περιγράφουμε συστηματικές μελέτες σχετικά με την αλληλεπίδραση παθητικοποιημένων από συνδέτη HfO2, ZrO2 και Hfx Zr1−x O2 NPs με το χουμικό οξύ του ποταμού Suwannee (HA) ως μοντέλο για το NOM. Τα ακόλουθα θέματα έχουν εξεταστεί σε αυτή την εργασία: (1) ο καταμερισμός υδρόφοβων επικαλυμμένων MO-NPs, με ή χωρίς HA, στην υδατική φάση. (2) εξέταση αλληλεπιδράσεων που επιτρέπουν στο ΗΑ να σταθεροποιεί κολλοειδώς τα MO-NP (δηλαδή, ηλεκτροστατικές, συντονιστικές και διασκορπιστικές αλληλεπιδράσεις). και (3) τη σημασία τόσο της επιφάνειας NP (κρυσταλλική δομή: μονοκλινική ή τετραγωνική) όσο και των παθητικοποιητικών προσδεμάτων (επιφανειακή επίστρωση: οξείδιο τρι-ν-οκτυλοφωσφίνης (TOPO)) στη σταθεροποίηση των NPs από ΗΑ στο νερό. Το TOPO είναι ένας πανταχού παρών υποκαταστάτης στη νανοεπιστήμη και χρησιμοποιείται συνήθως σε θερμή κολλοειδή σύνθεση. Αυτή η μελέτη παρέχει μια μηχανιστική αναφορά της σταθεροποίησης σε υδατική φάση διαφορετικών τύπων υδροφοβικά επικαλυμμένων MO-NPs με και χωρίς HA. Συγκεκριμένα, επιχειρήσαμε να χαρακτηρίσουμε άμεσα διεργασίες που διέπουν την προσρόφηση και συσσωμάτωση των MO-NPs με HA.

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συντέθηκαν με τη μη υδρολυτική συμπύκνωση κολλοειδούς-πηκτώματος αλκοξειδίων μετάλλων με αλογονίδια μετάλλων χρησιμοποιώντας το TOPO (Strem Chemicals, MA, USA) ως συνδετήρα συντονισμού [11]. Αυτή η συνθετική προσέγγιση παρέχει μονοδιασπορά και εξαιρετικό έλεγχο της κρυσταλλικής δομής, μεγέθους και στοιχειομετρίας. Ο συνδέτης TOPO συντεταγίζεται σε επιφανειακά άτομα, ολοκληρώνοντας το κέλυφος συντονισμού για υποσυντονισμένες μεταλλικές τοποθεσίες και επομένως χρησιμεύει ως παθητικοποιητική επίστρωση. Η χρήση συστημάτων μονοδιασποράς παρέχει τυποποίηση που απαιτείται για προσεκτικές μηχανιστικές μελέτες και αποκλείει τη συσκότιση από πολυδιασπορές στο μέγεθος των σωματιδίων, την επιφανειακή κάλυψη και την κρυσταλλική δομή, κάτι που θα περιέπλεξε σημαντικά τις μελέτες των MO-NPs με NOM. Όλες οι σκόνες ΜΟ-ΝΡ ήταν εύκολα διασπειρόμενες σε εξάνιο. Παρασκευάστηκαν και χρησιμοποιήθηκαν περίπου 500–1000 mg/L εναιωρήματα NP για τα πειράματα μεταφοράς φάσης. Αυτές οι συγκεντρώσεις επιλέχθηκαν για ευκολία στην ανίχνευση NPs. Δεν έχει ακόμη προκύψει συναίνεση σχετικά με το τι συνιστά περιβαλλοντικά ρεαλιστική συγκέντρωση MO NP. Η χρήση των εν λόγω συγκεντρώσεων μας επιτρέπει να αναπτύξουμε τυπικά εργαλεία μικροσκοπίας και φασματοσκοπίας για την αποσαφήνιση της φύσης των αλληλεπιδράσεων NP-NOM και πράγματι τέτοιες αλληλεπιδράσεις είναι πιθανό να διατηρηθούν ακόμη και σε χαμηλές συγκεντρώσεις. Για δοκιμές υδατικής διαλυτότητας, σκόνες NP διασπείρονται σε νερό. Το πρότυπο Suwannee River HA (SRHA-II) αγοράστηκε από την International Humic Substances Society (St. Paul, MN, ΗΠΑ). Η χρήση του καλά χαρακτηρισμένου SRHA-II επιτρέπει επίσης την τυποποίηση, η οποία διατυπώνεται σε διάφορα διεθνή εργαστήρια ως επείγον στόχος για τη δημιουργία γενικεύσιμων μέσων αξιολόγησης της τύχης και της μεταφοράς του ENM. Δεδομένου του πρωταρχικού μας στόχου να διασαφηνίσουμε τη μηχανιστική βάση για τη μεταφορά φάσης ENM, η χρήση καλά χαρακτηρισμένων NOM αποκτά ύψιστη σημασία. Για την ανάλυση Hf/Zr, τα πρότυπα μετάλλων (διαλυτά σε φθοριούχα μέταλλα) και τα συμπυκνωμένα HF και HNO3 βαθμού Aristar Ultra από την BDH Chemicals (West Chester, PA, ΗΠΑ) χρησιμοποιήθηκαν σε τυπική παρασκευή και πέψη με οξύ, αντίστοιχα. Απιονισμένο νερό (DI) από ένα σύστημα νερού Barnstead NANOpure (ΗΠΑ) χρησιμοποιήθηκε για την παρασκευή όλων των υδατικών διαλυμάτων ( ειδική αντίσταση = 18,2 M/cm). 2.2. Πειράματα μεταφοράς φάσης Ένα κλάσμα 5-mL του εναιωρήματος ΜΟ-ΝΡ (σε εξάνιο) αναμίχθηκε με 5 mL 20 mg/L ΗΑ σε DI νερό (ρΗ ~ 4,4) σε ένα διαυγές φιαλίδιο. Αυτή η πειραματική κατασκευή αναφέρεται ως «διάταξη μεταφοράς φάσης». Ένα διάλυμα HA 20 mg/L περιέχει 12,5 mg/L διαλυμένου οργανικού άνθρακα που είναι εντός των τυπικών επιπέδων των φυσικών νερών (0,1–200 mg/L) [33]. Το χαμηλό φυσικό pH που χρησιμοποιείται εδώ είναι επίσης αντιπροσωπευτικό του χαμηλού pH του ποταμού Suwanee όπου έγινε δειγματοληψία του ΗΑ. Παρόμοιες συνθήκες χρησιμοποιήθηκαν στην προηγούμενη εργασία μας για τα CdSe QDs [30,31]. Οι ρυθμίσεις μεταφοράς φάσης προετοιμάστηκαν επίσης σε νερό DI (χωρίς ΥΑ) για να χρησιμεύσουν ως έλεγχοι. Πρόθεσή μας ήταν να μελετήσουμε τις αλληλεπιδράσεις των MO-NPs με ανάλογα πραγματικών περιβαλλοντικών δειγμάτων, που περιέχουν ελεγχόμενες συγκεντρώσεις ΗΑ. Μεταξύ των μετρήσεων, κάθε διάταξη προστατεύτηκε από το φως χρησιμοποιώντας αλουμινόχαρτο και αναδεύτηκε συνεχώς σε θερμοκρασία δωματίου χρησιμοποιώντας αναδευτήρα πλατφόρμας ταλάντωσης για να διεγείρει τις διαδικασίες φυσικής ανάμειξης και διάχυσης υγρών. Η ανάμιξη πραγματοποιήθηκε για 15 ημέρες (~2 εβδομάδες). Για αναλύσεις μετάλλων, προετοιμάστηκαν μεμονωμένες ρυθμίσεις μεταφοράς φάσης (0, 1, 3, 5, 10 και 15 ημέρες). Αυτό αποφεύγει τα σφάλματα δειγματοληψίας λόγω της διεπιφανειακής συσσωμάτωσης (δηλαδή, αφαίρεση κροκιδωμένων συσσωματωμάτων καθώς μειώνεται ο όγκος του διαλύματος).

2. Experimental

2.3. Analysis and tool

2.1. Materials

Qualitative and quantitative analyzes were performed using ␨ potential measurements, dynamic light scattering (DLS), Raman and Fourier transform infrared (FTIR) spectroscopy, and inductively coupled plasma mass spectrometry (ICP-MS).

Monoclinic (m-) HfO2 and Hf0.37 Zr0.63 O2 and tetragonal (t) ZrO2 and Hf0.37 Zr0.63 O2 NP powders used in these experiments

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Fig. 1. HRTEM images of (A1–2) m-HfO2 and t-ZrO2 NPs and (B3–4) MO-NPs with HA transfer in aqueous solution. Insets (i–ii) highlight the dominant crystal planes of the NPs. Lattice spacings are assigned based on monoclinic (JCPDS# 780050) and tetragonal (JCPDS# 881997) structures.

␨-Potential and DLS data were collected using a Zetasizer Nano ZS90 instrument (Malvern Instruments, Malvern Hills, UK). TEM measurements were performed using a JEOL JEM-2010 (Tokyo, Japan) operating at an accelerating voltage of 200 kV. Raman spectra were obtained at room temperature using a Horiba JobinYvon (Villeneuve d'Ascq, France) Labram HR Raman spectrometer using 784.51 nm laser excitation from a diode laser. FTIR spectra were collected using a Nicolet-Magna (USA) 550 dry air purged spectrometer with a spectral resolution of 4 cm-1. The total concentrations of Hf and Zr present in the organic and aqueous phases were quantified by ICP-MS. ICP-MS measurements were performed with a Thermo Scientic (Germany) X-Series 2 instrument. The concentrations of Hf and/or Zr in the samples were determined using an external calibration curve. Details of the sampling and acid digestion protocol are described in the Supporting Information (SI).

3. Results and discussion 3.1. Characteristics of MO-NPs Fig. Figures 1A and S1A show grating-resolution TEM images and high-resolution TEM (HRTEM) images of m-HfO2, t-ZrO2, mHf0.37Zr0.67O2, and t-Hf0.37Zr0.67O2 NPs. All NP surfaces are passivated with TOPO and phosphonate ligands. The m-HfO2 particles have an elongated rice grain-like morphology with aspect ratios ranging from 3 to 4. In contrast, the t-ZrO2 NPs have a quasi-spherical morphology with an average diameter of 3.3 nm. NP m-Hf0.37 Zr0.67 O2 solid solutions are slightly elongated, while NP t-Hf0.37 Zr0.67 O2 are almost spherical [11]. Representative XRD patterns of MO-NPs used in phase transfer experiments are shown in the figure. S2. figs. Figures 1A and S1A show the monodispersity of the NPs used. HRTEM images show

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τις εκτεθειμένες κρυσταλλικές όψεις για κάθε σύνολο σωματιδίων. Τα μονοκλινικά NP δείχνουν σταθερά μια προτίμηση για {1 0 0} και {1 1 1} κρυσταλλικές όψεις, ενώ τα τετραγωνικά NP επιδεικνύουν κατά κύριο λόγο {1 0 1} επιφανειακά τερματικά επίπεδα. Αυτές οι εκχωρήσεις είναι συνεπείς με υπολογισμούς επιφανειακής ενέργειας για m-HfO2 και t-ZrO2, οι οποίοι υποδεικνύουν προνομιακή ενεργειακή σταθεροποίηση για αυτά τα επίπεδα στις μονοκλινικές και τετραγωνικές κρυσταλλικές δομές [34,35]. Δεδομένης της υδρόφοβης φύσης του TOPO, θα ήταν λογικό να αναμένουμε την πρώτη προσέγγιση ότι τέτοια NPs θα έχουν ασήμαντες δυνατότητες διασποράς ή κινητικότητα στο υδάτινο περιβάλλον. Πράγματι, αυτά τα NP εμφανίζουν πολύ χαμηλές υδατικές διαλυτότητες: 2 εβδομάδες. Απουσία ΗΑ, η μεταφορά φάσης λαμβάνει χώρα με σχετικά βραδύτερο ρυθμό και τα μεταφερόμενα MONP τείνουν να κροκιδώνονται και τελικά να καθιζάνουν στον πυθμένα του φιαλιδίου. Πράγματι, οι εικόνες TEM, τα φασματικά δεδομένα ␨-δυναμικού, DLS και FTIR υποδηλώνουν άμεση αλληλεπίδραση μεταξύ των MO-NP και του HA. Τρεις διακριτές αλληλεπιδράσεις του ΗΑ με τα MO-NPs που οδηγούν σε διασπορά και σταθεροποίηση υδατικής φάσης μπορούν να προβλεφθούν [30,31,45]. Η πρώτη διαδικασία περιλαμβάνει την επικάλυψη των MO-NPs με HA μέσω μη ειδικής προσρόφησης χουμικών στο κέλυφος του συνδέτη TOPO. Τα αμφίφιλα χαρακτηριστικά του ΗΑ επιτρέπουν τον σχηματισμό ψευδομικελικών συσσωματωμάτων που επικαλύπτουν τα NPs, όπου οι υδρόφοβες αρωματικές και ετεροαλειφατικές περιοχές σχηματίζουν μια υδρόφοβη εσωτερική κοιλότητα, ενώ τα κρεμαστά υδρόφιλα καρβοξυλικά τμήματα, φαινολικά και αμίνης κατευθύνονται προς τα έξω, σταθεροποιώντας την ηλεκτροσταθερότητα στο νερό. [38,46]. Υπό την προϋπόθεση ότι οι επιφανειακοί συνδέτες NP εμπλέκονται σε αυτή την αλληλεπίδραση, αυτός ο μηχανισμός δεν αναμένεται να δείξει καμία διάκριση μεταξύ μονοκλινικών και τετραγωνικών κρυσταλλικών δομών. Η δεύτερη διεργασία περιλαμβάνει δοτικές αλληλεπιδράσεις μεταξύ των επιφανειών ΗΑ και ΜΟ, όπου το καρβοξυλικό οξύ (εκτιμάται ότι είναι 10% στο SRHA-II, [47]), τα φαινολικά και τα τμήματα αμίνης που είναι άφθονα στη δομή του ΗΑ χρησιμεύουν ως ευέλικτη χηλική πολυοδοντωτή ένωση συνδέτες [38,48-51]. Η αλληλεπίδραση που περιλαμβάνει την επιφάνεια του NP είναι κατανοητή με δεδομένες τις εξαιρετικά αντιδραστικές επιφάνειες των υλικών νανοκλίμακας, όπου τα περισσότερα από τα συστατικά άτομα βρίσκονται στην επιφάνεια ή κοντά στην επιφάνεια [27,28,52]. Αυτός ο μηχανισμός βασίζεται στη διαθεσιμότητα και την προσβασιμότητα μεταλλικών θέσεων στην επιφάνεια MO-NP που μπορούν να συμμετέχουν σε συντονιστικές αλληλεπιδράσεις. Διαφορετικές κρυσταλλικές δομές έχουν διακριτικά επίπεδα, γωνίες και άκρες εκτεθειμένες στην επιφάνεια (Εικ. 1). Ως εκ τούτου, σε αντίθεση με τον μηχανισμό επικάλυψης, αυτή η αλληλεπίδραση μπορεί ενδεχομένως να παρέχει κάποια διάκριση μεταξύ μονοκλινικών και τετραγωνικών κρυσταλλικών δομών. Η τρίτη διαδικασία περιλαμβάνει ηλεκτροστατικές αλληλεπιδράσεις μεταξύ των συντονιστικά ακόρεστων επιφανειακών θέσεων στα MO-NPs (θετικά και αρνητικά φορτισμένα) και των καρβοξυλικών και φαινολικών τμημάτων (αρνητικά φορτισμένα) στη δομή ΗΑ. Αυτές οι διεργασίες (μικκυλιακή/επικάλυψη, συντονιστικές/υποκαταστατικές και ηλεκτροστατικές αλληλεπιδράσεις) φαίνεται να λειτουργούν συνεργιστικά για να διευκολύνουν τη μεταφορά φάσης και την επακόλουθη διασπορά σε υδατική φάση των μονοκλινικών MO-NP που παθητικοποιούνται με TOPO αλλά όχι με τα τετραγωνικά NP. Οι διαφορές στη συμπεριφορά μεταφοράς φάσης που παρατηρήθηκαν μεταξύ των δύο διαφορετικών πολύμορφων υποδηλώνουν ότι μεσολαβούν ειδικές και όχι μη ειδικές αλληλεπιδράσεις

μεταφορά φάσης? Σε αυτή την περίπτωση, οι συντεταγμένες/υποκατάστατες αλληλεπιδράσεις που εξαρτώνται από την επιφάνεια της δομής είναι η πιθανή γένεση της διακριτής αντιδραστικότητας. Κατά την ανάμιξη με Η2Ο/ΗΑ, ορισμένοι από τους συνδέτες TOPO είναι πιθανό να εκτοπιστούν. Τα φάσματα FTIR των συσσωματωμάτων MO-HA που μεταφέρονται φάσης υποδηλώνουν ότι η ποσότητα του TOPO μειώνεται σε σύγκριση με τα επικαλυμμένα με TOPO NPs. Η αφαίρεση του TOPO παρέχει πρόσβαση σε οξόφιλες, συντονιστικά ακόρεστες κατιονικές θέσεις στις επιφάνειες MO-NP στις οποίες μπορούν να προσπελαστούν τα τμήματα καρβοξυλικού οξέος του ΗΑ (οι ηλεκτροστατικές αλληλεπιδράσεις πιθανότατα προκαλούν την αρχική προσέγγιση των τμημάτων HA και NPs). Εκτός από τη μετατόπιση του προσδέματος, η ατελής κάλυψη/παθητικοποίηση των αρχικών NPs από το TOPO θα καθιστούσε επίσης τις επιφανειακές θέσεις Hf/Zr διαθέσιμες για δέσμευση με HA. Οι ακαθαρσίες στο TOPO τεχνικής ποιότητας (90%), ιδιαίτερα τα αλκυλοφωσφονικά και αλκυλοφωσφινικά οξέα που έχει αποδειχθεί ότι παίζουν ενεργό ρόλο στην παθητικοποίηση επιφανειών κβαντικών κουκκίδων, ράβδων και συρμάτων CdSe [53-56], θα μπορούσαν επίσης να επηρεάσουν τις αλληλεπιδράσεις μεταξύ ΗΑ και την επιφάνεια NP [31]. Η κλασματική επιφανειακή κάλυψη των ομάδων επικάλυψης και η ακριβής φύση του κελύφους παθητικοποίησης συνδέτη είναι πέρα από το πεδίο αυτής της μελέτης (τα κελύφη παθητικοποίησης συνδέτη παραμένουν να χαρακτηριστούν επαρκώς ακόμη και για κβαντικές κουκκίδες CdSe που είναι πιθανώς οι πιο ώριμες από αυτήν την κατηγορία υλικών) . Ωστόσο, όπως προτείνεται από το FTIR, ο σχηματισμός δεσμών μετάλλου-humate συνδέει τα MO-NPs με τα χουμικά κολλοειδή, τα οποία πιθανότατα έλκουν τα NPs στη διεπιφάνεια εξανίου/νερού έτσι ώστε τόσο τα υδρόφοβα NP όσο και τα υδρόφιλα κολλοειδή HA να μπορούν να επιδιαλυτωθούν επαρκώς . Στη συνέχεια, όπως περιγράφεται στη βιβλιογραφία [46,49,57], το εύκαμπτο χουμικό κολλοειδές μπορεί να υποβληθεί σε μοριακή αναδιάταξη και διασταύρωση με εγγύς τμήματα ΗΑ στη διεπιφάνεια εξανίου/νερού για να βελτιστοποιηθούν οι υδρόφοβες αλληλεπιδράσεις με τις κρεμαστά αλειφατικές αλυσίδες στον συνδέτη TOPO [41-43,58]. Επιπλέον, τα αποτελέσματά μας δείχνουν ότι το H2O από μόνο του επιτρέπει κάποια μεταφορά φάσης των NPs. Η οξοφιλικότητα των επιφανειών του πρώιμου οξειδίου μετάλλου μετάπτωσης μπορεί επίσης να επιτρέψει την εύκολη υποκατάσταση προσδέματος από Η2Ο, η οποία μπορεί τελικά να οδηγήσει σε αξιόλογη υδροξυλίωση των επιφανειών ΜΟ. Οι συγγένειες διαφορετικών κρυσταλλογραφικών όψεων για το ΗΑ μπορεί εύλογα να θεωρηθεί ότι είναι παράλληλες με την πιθανότητα υποκατάστασης συνδέτη από Η2Ο, γεγονός που μπορεί να εξηγήσει τη σημαντική μεταφορά φάσης που παρατηρείται για τα μονοκλινικά NPs ακόμη και απουσία ΗΑ. Η μεταφορά φάσης δείχνει διαφορές μεταξύ Η2Ο και ΗΑ μόνο σε μικρότερες χρονικές κλίμακες, όταν η αλληλεπίδραση MO-H2O είναι πιθανώς περιορισμένη από την υποκατάσταση συνδέτη. Η ενδοεπιφανειακή θολότητα και η βραχύχρονη μεταφορά φάσης που σημειώνονται στα δείγματα ελέγχου απουσία ΗΑ πιθανώς προκύπτουν από τη μετατόπιση ορισμένων προσδεμάτων TOPO από Η2Ο. Συνάδει με τον προτεινόμενο μηχανισμό, καθώς τα μόρια Η2Ο που συντονίζονται στην επιφάνεια δεν έχουν τα αμφιφιλικά χαρακτηριστικά του ΗΑ , δεν είναι σε θέση να σταθεροποιήσουν επαρκώς τα επικαλυμμένα με TOPO MO-NPs στην υδατική φάση. Με άλλα λόγια, η υποκατάσταση συνδέτη του TOPO από Η2Ο μπορεί να προκαλέσει καθίζηση στη διεπιφάνεια εξανίου/νερού αλλά δεν επιτρέπει την κολλοειδή σταθεροποίηση στην υδατική φάση απουσία ΗΑ. Η προτιμώμενη δέσμευση ΗΑ ή Η2Ο σε μονοκλινικές αντί για τετραγωνικές επιφάνειες μπορεί να αποτελέσει τη βάση για την παρατηρούμενη επιλεκτικότητα μεταφοράς φάσης όπως μεταξύ m-HfO2 και t-ZrO2 NPs και μεταξύ m-Hf0,37 Zr0,63 O2 και t- Hf0,37 Zr0,63 O2. Όπως φαίνεται στο Σχ. 1Α, τα NP m-HfO2 (και m-Hf0,37 Zr0,63 O2 ) εκθέτουν κατά προτίμηση επιφάνειες {1 0 0} και {1 1 1}, ενώ για t-ZrO2 (και tHf0,37 Zr0 .63 O2 ) Οι επιφάνειες NPs {1 0 1} προτιμώνται ενεργειακά. Ενώ η διαφορά μεταξύ m-HfO2 και t-ZrO2 NPs όσον αφορά τη μεταφορά φάσης θα μπορούσε να σχετίζεται με την έκταση/ισχύ των δεσμών Hf-HA έναντι των δεσμών Zr-HA, προκύπτει από το m-Hf0,37 Zr0,63 O2 και t-Hf0,37 Zr0,63 O2 NPs (ίδια χημική σύνθεση) υποδηλώνουν ότι η μεταφορά φάσης ανταποκρίνεται περισσότερο στις αλλαγές στην κρυσταλλική δομή. Κατά συνέπεια, παρά το γεγονός ότι και οι δύο έχουν δυνητικά αντιδραστικές επιφάνειες, εικάζουμε ότι οι διαφορές στη μεταφορά φάσης των NPs μπορεί να προέρχονται από (α) τις σχετικές συγγένειες δέσμευσης των διαφορετικών επιφανειών σε μονοκλινικά και τετραγωνικά NP για TOPO και HA. (β) ο βαθμός του

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συντονιστικός ακόρεστος και στερεοχημική παρεμπόδιση για θέσεις μετάλλων μετάπτωσης στα διαφορετικά επίπεδα επιφάνειας. και (γ) την πυκνότητα των εκτεθειμένων θέσεων κατιόντων μετάλλου μεταπτώσεως. Ένας συνδυασμός αυτών των παραγόντων θα μπορούσε να καταστήσει τις μεταλλικές θέσεις λιγότερο διαθέσιμες και τη μετατόπιση των προσδεμάτων TOPO από λειτουργίες HA πιο δύσκολη για τα τετραγωνικά NPs. Ελλείψει δεσμών μετάλλου-humate, η μεταφορά φάσης μπορεί να μην ξεκινήσει τόσο εύκολα, με αποτέλεσμα τις χαμηλές αποδόσεις μεταφοράς φάσης που παρατηρούνται για τα t-ZrO2 και t-Hf0,37 Zr0,63 O2 NPs. Ανάλογη προτίμηση για δέσμευση διαφορετικών επιφανειών αναμένεται επίσης για συντονιστικές αλληλεπιδράσεις με μόρια Η2Ο. Ο υπολογισμός των συνάφειων δέσμευσης, της έκτασης του επιφανειακού ακόρεστου και της πυκνότητας κατιόντων και ανιόντων στην επιφάνεια του NP ξεφεύγει από το πεδίο εφαρμογής αυτής της μελέτης. Αυτές οι μετρήσεις δεν έχουν πράγματι επικυρωθεί πειραματικά για συστήματα κολλοειδούς NP παθητικοποιημένα από συνδέτη με οποιοδήποτε βαθμό ακρίβειας. Ωστόσο, τα εκτεταμένα δεδομένα χαρακτηρισμού μας είναι επαρκή για την εξαγωγή συμπερασμάτων σχετικά με τους μηχανισμούς που υπαγορεύουν τη μεταφορά φάσης και τη σταθεροποίηση υδατικής φάσης αυτών των NP. 4. Συμπεράσματα Οι αλληλεπιδράσεις μεταξύ HA και MO-NPs υποδεικνύουν το ρόλο του ΗΑ τόσο ως συντονιστικού συνδετήρα όσο και ως αμφίφιλου επιφανειοδραστικού. Τα αποτελέσματά μας παρέχουν περαιτέρω πειραματική επικύρωση των θεωρητικών προβλέψεων [14] και των πειραματικών παρατηρήσεων [15-17] των τροποποιήσεων στην κολλοειδή σταθερότητα των MO-NPs κατά την απόκτηση επικαλύψεων NOM. Σε αυτή τη μελέτη, το HA και το H2O παρουσιάζουν μια ξεχωριστή προτίμηση για μονοκλινικά παρά τετραγωνικά MONP, πιθανώς λόγω ισχυρότερων δεσμευτικών συνάφειων με μονοκλινικές επιφάνειες (συντονιστικές/υποκαταστατικές αλληλεπιδράσεις) και της ευκολίας σχηματισμού κυλινδρικών ψευδο-μικκυττάρων ΑΛΛΗΛΕΠΙΔΡΑΣΗ). Η έκταση της μεταφοράς φάσης και ο βαθμός κολλοειδούς σταθεροποίησης που παρατηρείται για καλά καθορισμένα MO-NP με υδρόφοβες επικαλύψεις παρουσία ΗΑ υπογραμμίζουν επίσης τη σημασία της ανάπτυξης μιας λεπτομερούς κατανόησης των πιθανών περιβαλλοντικών μετασχηματισμών των ENP. Η χαρακτηριστική επιλεκτικότητα στη μεταφορά φάσης που προκαλείται από ΗΑ και τη διασπορά διαφορετικών πολυμορφών υποδηλώνει ότι οι αλληλεπιδράσεις διαφορετικών ανόργανων ENM με το NOM, που υπαγορεύουν την έκταση της μεταφοράς και σταθεροποίησης NP, δεν είναι απαραίτητα γενικεύσιμες και ότι μπορεί να είναι δυνατός ο σχεδιασμός NM για την ελαχιστοποίηση χρόνο παραμονής τους στο υδάτινο περιβάλλον. Από αυτή την άποψη, απαιτείται περαιτέρω έρευνα για τον προσδιορισμό των πραγματικών συγγενειών δέσμευσης του ΗΑ σε μονοκλινικές και τετραγωνικές επιφάνειες και για τη διερεύνηση της επίδρασης διαφορετικών επικαλύψεων επιφανειών. Ευχαριστίες Αυτή η εργασία υποστηρίχθηκε κυρίως από την Υπηρεσία Προστασίας του Περιβάλλοντος των ΗΠΑ (Grant# R833861). Η SB αναγνωρίζει τη μερική υποστήριξη αυτής της εργασίας από το Εθνικό Ίδρυμα Επιστημών με DMR 0847169. Αναγνωρίζουμε το Πρόγραμμα MRI NSF CHE 0959565 για την απόκτηση του οργάνου ICP-MS. Τόσο η CSIRO όσο και η ΥΠΠ των ΗΠΑ δεν έχουν υποβάλει αυτό το χειρόγραφο σε εσωτερική αναθεώρηση ομοτίμων και πολιτικής. Ως εκ τούτου, δεν πρέπει να συναχθεί καμία επίσημη έγκριση. Παράρτημα A. Συμπληρωματικά δεδομένα Συμπληρωματικά δεδομένα που σχετίζονται με αυτό το άρθρο βρίσκονται στην ηλεκτρονική έκδοση στη διεύθυνση doi:10.1016/j.jhazmat.2011.09.028. Αναφορές [1] P.J.J. Alvarez, V.L. Colvin, J. Lead, V. Stone, Προτεραιότητες έρευνας για την προώθηση της οικο-υπεύθυνης νανοτεχνολογίας, ACS Nano 3 (2009) 1616–1619. [2] A. Maynard, R.J. Aitken, T. Butz, V. Colvin, K. Donaldson, G. Oberdorster, M.A. Philbert, J. Ryan, A. Seaton, V. Stone, S.S. Tinkle, L. Tran, N.J. Walker, D.B. Warheit, Safe handling of nanotechnology, Nature 444 (2006) 267–269.

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Journal of Hazardous Materials 196 (2011) 311-317

Content is available on SciVerse ScienceDirect

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Reduction of As(V) to As(III) with commercial ZVI or As(0) with acid-treated ZVI Fenglong Sun a, Kwadwo A. Osseo-Asare b, Yongsheng Chen c, Brian A. Dempsey a,∗ a

Department of Civil and Environmental Engineering, Penn State University, UP, USA Department of Materials Science and Engineering, Penn State University, UP, USA c Department of Energy and Mineral Engineering, Penn State University, UP, USA b

article

i n f o

Article history: Received April 24, 2011 Received in revised form September 7, 2011 Accepted September 8, 2011 Available online September 14, 2011 Keywords: Zerovalent iron-arsenic reduction X-ray absorption spectroscopy (XAS)

a b s t r a c t Zero Iron (ZVI) consists of an elemental iron core surrounded by a shell of corrosion products, especially magnetite. ZVI is used for in situ removal or immobilization of various contaminants, but the mechanisms of arsenic removal remain controversial, and the mobility of arsenic after reaction with ZVI is uncertain. These questions were addressed by separately studying the reactions of As(V) with magnetite, commercial ZVI, and acid-treated ZVI. Strictly anoxic conditions were used. Adsorption of As(V) on magnetite was rapid with a pH dependence similar to previous reports using acidic conditions. As(V) is not reduced by magnetite and Fe(II), although the reaction is thermodynamically spontaneous. As(V) reactions with ZVI were also fast and no lag phase was observed, which was in contrast to previous reports. Commercial ZVI reduced As(V) to As(III) only when As(V) was adsorbed, ie at pH < 7. As(III) was not released into the solution. Acid-treated ZVI reduced As(V) to As(0), as shown by liquid chemical analyzes and XANES/EXAFS. Comparisons were made between the reactivity of acid-treated ZVI and nano-ZVI. If true, acid-treated ZVI may offer similar reactive advantages at a lower cost. © 2011 Elsevier B.V. All rights reserved.

1. Introduction: High concentrations of arsenic in groundwater have been reported worldwide, especially in Bangladesh and Taiwan [1]. Arsenic in the aquatic environment is usually found in inorganic forms such as arsenic (As(V)) and arsenite (As(III)). As(III) is usually more mobile and toxic than As(V), although As(III) can also be immobilized in the presence of sulfide. Elevated arsenic concentrations in groundwater may occur due to reductive dissolution of iron oxide sorbents and subsequent reduction and mobilization of As(III), desorption of As(V) under alkaline pH conditions, especially in the presence of phosphates or other competitive adsorbents or oxidizing sulfide materials [1, 2]. Zevalent iron (ZVI) is used to remove organic and inorganic contaminants, including chlorinated solvents, nitrates, uranyl ions, chromate, lead, and arsenic [3,4]. ZVI can be incorporated into permeable reactive dams or nano-ZVI (nZVI) can be injected into contaminated soil [5]. ZVI is also found in some drinking water treatment systems [6]. ZVI is usually stated to have a core-shell structure. The shell contains oxidized iron, which is mainly magnetite and often with maghemite (␥-Fe2 O3 ) or lepidocrocite (␥-FeOOH) [3,7-10].

∗ Corresponding author. Phone: +1 814 865 1226; fax: +1 814 863 7304. E-mail:[email protected](B.A. Dempsey). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.029

ZVI immobilizes arsenic by adsorption of As(V) or As(III) to iron corrosion products in the shell surrounding the elemental iron core, and this is sometimes accompanied by the reduction of As(V) to As(III) [11-14] . Detailed mechanisms for the removal of many pollutants by ZVI have been reported. Fe(0), dissolved Fe(II), solid Fe(II) and H2 have been proposed as elemental reductants [3,9,15]. However, the mechanisms of arsenic removal, especially related to redox reactions, are still controversial. The rate and extent of As(V) reduction by ZVI may depend on the experimental conditions. In different studies, 25% of initial As(V) was reduced to As(III) by nano-ZVI at neutral pH after 90 days [13,16], As(V) was partially reduced to As(III) by commercial ZVI at slightly basic pH after 60 days [17] and there was no reduction of As(V) by iron wires [18]. It was also reported that As(III) was reduced to As(0) by acid-treated iron filings [19]. Magnetite is often considered the dominant component in the weathered ZVI shell. In an attempt to identify the mechanisms by which ZVI immobilized contaminants, Lago et al. [20,21] used mechanical milling to produce a magnetite/ZVI reagent that reduces methylene blue, H2O2, and Cr(VI). Other iron oxides (␣Fe2 O3, FeOOH or ␥-Fe2 O3) mixed with ZVI were much less reactive. It has been suggested that the semiconducting behavior of magnetite is important for effective pollution reduction. The reactivity of magnetite may depend on whether the ZVI has been exposed to air. In this context, White and Peterson [22] showed that magnetite reduces Cr(VI) at a much faster rate under anoxic conditions than under oxic conditions. It also proved to be so

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"Stoichiometric magnetite" produced and preserved under strictly anoxic conditions has a Fe(II)/Fe(III) ratio close to 0.5 and is a stronger reducing agent than "non-stoichiometric magnetite" produced or stored in the presence of O2, resulting in an Fe(II) ratio of /Fe(III) k( FeAsO4 2− ) which is consistent with reduced adsorption of As(V) with increasing pH. As (V) adsorption on magnetite creates similar inner sphere complexes as on HFO and goethite [42], and similar adsorption sources have been reported [24-26,42]. As(V) adsorption on magnetite under anoxic conditions in this study showed a similar pH dependence as previously reported for acidic conditions. All dissolved and adsorbed arsenic in magnetite experiments was recovered as As(V), even when the experiment was performed in the presence of excess dissolved Fe(II). Data on the oxidation state of arsenic are presented in Fig. S1. Although we used stoichiometric magnetite (a stronger reducing agent than non-stoichiometric magnetite [22,23]), strictly anoxic

Fig. 5. Adsorption of As(V) and reduction to As(III) using commercial ZVI (2.0 g/0.02 L) in an anoxic environment: (a) pH = 6.5 ± 0.5; (b) pH = 8.5 ± 0.5; (c) pH = 10.

environment and stoichiometric excess of dissolved Fe(II), magnetite did not reduce As(V) at pH 5-10. 3.3. Reactions of As(V) with commercial ZVI The results of experiments with As(V) and commercial ZVI at pH 6.5 ± 0.5, 8.5 ± 0.5 and 10 are shown in Figure 5, where the concentrations of dissolved and adsorbed As (V) and As(III) are plotted against time. There was a partial reduction of As(V) to As(III) in the two lower pH ranges and a certain loss of As(V) + As(III) at pH 8.5 ± 0.5.

Οι F. Sun et al. / Journal of Hazardous Materials 196 (2011) 311–317

Fig. 6. Adsorption of As(V) and reduction to As(0) using acid-treated ZVI (2 g in 0.02 L) in an anoxic environment at pH 7. No As(III) was produced.

t1/2 for adsorption and reduction in As(III) was Mn2+ (Table 4). This implies that the negative charge of the birnessite layer is mostly balanced by H+ and K+. Mn2+ was not detected during Pb2+ adsorption by HB, indicating the presence of small amounts of Mn2+/3+ in the interlayer, which was consistent with XPS results. The amount of released H+ and K+ was positively correlated with the adsorption capacity of Pb2+, but the amount of Co2+ and Mn2+ increased with increasing cobalt content. As the cobalt content increases, more Mn2+ ions are released into the solution. These low-valent Mn ions were formed by the reduction of Mn3+/4+ with Co2+. About 9.6%, 7.4%, 7.9% and 13.1% of the total cobalt was released in HC2, HC5, HC10 and HC20, respectively. It showed that there may be some Co2+ in the interlayer or surface of birnessite in addition to Co3+ when the initial Co2+ concentration was high [21]. Co2+ can be replaced by Pb2+ during adsorption. In addition, Co3+ in the interlayer can also be driven by Pb2+ in solution in connection with redox reactions, but the underlying mechanisms are still unclear. Aqueous Pb2+ speciation is shown in Fig. 7 as a function of pH in the range 3-11 (calculated using ECOSAT4.9 [54]). When the pH was below 5.4, Pb2+ and [Pb(NO3)]+ were the major species. As it increased to 5.4, the formation of Pb(OH)2(s) began to limit the concentrations of aqueous species. At ~ pH 5.5, Pb(OH)2(s) was the major form present in the system. When the pH was above ~6.5, Pb(II) appeared mainly as Pb(OH)2(s). At pH 5.0, Pb(II) species existed in adsorption systems as Pb2+ (61.21%), [Pb(NO3)]+ (34.84%) and Pb(NO3)2 (3.95%). According to the law of charge conservation in the ion exchange process, after the adsorption of one mole of Pb2+, two moles of H+ and/or K+ or one mole of divalent cation (Me2+ ) are released, while the adsorption of [Pb(NO3 )] + will remove only one mole of H+ or K+ or 1/2 mole of Me2+ from the birnessite surface. This was also confirmed by the algebraic relationship between its sizes

n(Pb(II)) × (0.6121 × 2 + 0.3484) ≈ n(H+) + n(K+) + (n(Mn2+) + n(Co2+)) × 2 where n denotes the amount of adsorbed / released ion, and it is recorded in table 4. 3.4. The effects of Co2+ exchange on the oxidation of As(III)-arsenic pollution are a topic of great importance. Depending on the source, arsenic concentrations in natural water can vary up to several hundred milligrams per liter. Due to its acute toxicity to humans, the maximum contamination level (MCL) should be less than 10 ␮g L−1 for arsenic in drinking water [55]. Manganese oxides are active oxidants for the conversion of As(III) to As(V) under natural conditions. As prepared, birnessites were used to study the oxidative transformation of sodium arsenite on the surface of mineral water. The oxidation of arsenite was initially rapid, and the reaction rate then decreased to maintain equilibrium after 1–2 h. HC2 and HC5 had the same shape as HB. However, in the case of HC10 and HC20, the amount of As(III) oxidation gradually increased as the reaction progressed (Figure 8). The apparent oxidation capacity of As(III) in As(V) was calculated to estimate the oxidation capacity of birnessite, due to the fact that the adsorption and fixation of As(III) and As(V) took place simultaneously in the oxidation process [ 8 . ]. The calculated apparent oxidation capacity of As(III) by HB is 77.3% at equilibrium. After a 7-hour reaction, the conversion of As(III) to As(V) with HC2, HC5, HC10, and HC20 was 91.0%, 93.2%, 88.4%, and 60.2%, respectively. The high oxidation capacity towards As(III) was attributed to the participation of Co(III) in the reaction, since the typical reduction potential of Co3+ /Co2+ (E◦ = 1.92 V) is higher than MnO2 /Mn2+ (E◦ = 1.224 V) and Mn3+ /Mn2+ (E◦ = 1.5415 V) half-reaction [23]. This was already confirmed in our previous work using XPS analysis [24]. The relationship of As(III) concentration with time was analyzed by fitting the first-order rate equation to the segments 0–0.33 h of all five systems (Figure 9). Apparent reaction rate constants (kobs) for HB, HC2, HC5, HC10, and HC20 were calculated to be 0.0226, 0.0175, 0.0161, 0.0123, and 0.0035 min-1, respectively. The higher initial rate of As(III) oxidation reaction for HB than for Co-containing can be attributed to several reasons. First, oxygen atoms bound to Co3+ in inclusion birnessites will be retained more strongly than those bound to Mn3+/4+ due to the high crystal field stabilization energy (CFSE) of the low-spin Co3+ ion [15]. This would increase the activation energy at these sites, resulting in a slower reaction rate [56]. Second, As(III) oxidation of birnessite is a complex process. Research into the oxidation of arsenite by weakly crystalline manganese oxide showed that As(III) oxidation and As(V) sorption are strongly influenced by Mn AOS in

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Fig. 9. Linear regression analysis of normalized intake of As(III) obtained by birnessite.

structure of ␦-MnO2 [57]. It was expected that reactive Mn(III) sites on Mn oxide surfaces would be less reactive than Mn(IV) reactive sites with respect to As(III) oxidation [39,58,59]. Furthermore, XPS analysis and ion release during Pb2+ adsorption showed that certain amounts of Mn2+/3+ and Co2+/3+ were adsorbed on the birnessite surface. These low-valent ions also blocked reactive sites on the mineral surface. The release of Mn2+, Co2+ and K+ into the solution was also monitored. The concentrations of K+ released from HB, HC2, HC5, HC10 and HC20 during the arsenite oxidation process were 1998, 1523, 1247, 1068 and 669 mmol kg−1, respectively. However, Mn2+ and Co2+ were not detected when HB, HC2 and HC5 were used as oxidants. For HC10 and HC20, 13, respectively 0, 36 and 8 mmol kg−1 Mn2+ and Co2+ were determined. There are two reasons for this small release of Mn2+/Co2+: (i) The surface of MnO2 particles contained negatively charged surface functional groups (≡Mn–O−), so soluble Mn2+/3+ and Co2+/3+ were formed by reductive solution. Bearing birnessites were adsorbed are on the surface at pH 7.00 in this study and (ii) a precipitate of Co3 (AsO4)2 (pKsp = 28.17) was formed [60] and probable Krautite [61]. Oxidation of As(III) by manganese oxide has been an important reaction both in the natural cycle of As and in the development of remediation technology to reduce As(III) concentration in drinking water. In the presence of co-containing birnessite, As(III) in wastewater or groundwater will be oxidized to As(V). As(III) has higher mobility and weaker adsorption, and is therefore more toxic than As(V) [8]. Since As(V) exists as a deprotonated oxyanion over a wide pH range [23] and has a high affinity for mineral surfaces, the oxidation of As(III) to As(V) not only reduces its toxicity, but also facilitates the speciation of As. Metal compounds (Fe/Al oxides, hydroxides, etc.) are the most commonly used adsorbents for As, due to their higher removal efficiency at a lower cost compared to many other adsorbents [4]. The maximum amount of As(V) produced during the oxidation of As(III) with birnessite was significantly increased in the presence of goethite. The combined effects of oxidation (using birnessite) and adsorption (using goethite) led to rapid oxidation and immobilization of As and reduction of As toxicity in the environment [62]. Therefore, strong oxidation of arsenite with manganese oxides, followed by adsorption of arsenic with Fe/Al compounds as adsorbents, is a viable approach for the treatment of As(III)-contaminated water systems and deserves further investigation. 4. Conclusions The exchange of Co2+ ions with birnessite was carried out at different concentrations of Co2+. Co2+ ions are completely retained by birnessite at low concentrations. However, when the initial Co/Mn molar ratio was increased to 0.2, only 80% of Co2+ could be found in

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δομή. Η επαγωγή του Co2+ δεν είχε καμία επίδραση στην κρυσταλλική δομή και τη μορφολογία του μπιρνεσίτη. Η κρυσταλλικότητα των μπιρνεσιτών που περιέχουν Cocontaining μειώθηκε σταδιακά και αυξήθηκαν οι συγκεκριμένες επιφάνειες. Το σθένος του Co ήταν αποκλειστικά +3 και το Mn AOS των Co-containing birnessites σταδιακά μειώθηκε. Η περιεκτικότητα σε ομάδες υδροξυλίου στη δομή των Βιρνεσιτών που περιέχουν συν-περιέχοντα μειώθηκε σταδιακά, γεγονός που ευθύνεται για τις μειωμένες ικανότητες προσρόφησης Pb2+. Ενισχύθηκε η οξείδωση του As(III) από τον Βιρνεσίτη που περιέχει το ίδιο. Αντίθετα, με μια αύξηση στη συγκέντρωση κοβαλτίου, η σταθερά του αρχικού ρυθμού αντίδρασης μειώθηκε σημαντικά. Κατά τη διάρκεια της διαδικασίας οξείδωσης του Co2+, το Mn(IV) ήταν πιο πιθανό η καταβύθιση ηλεκτρονίων και στη συνέχεια μειώθηκε σε Mn(III). Η παρούσα εργασία παρέχει μια νέα εικόνα της περιβαλλοντικής χημικής συμπεριφοράς και του μηχανισμού αλληλεπίδρασης των οξειδίων του κοβαλτίου και του μαγγανίου. Επιπλέον, αυτά τα τροποποιημένα υλικά έχουν υψηλότερη ικανότητα προσρόφησης για το Pb2+ από πολλά άλλα προσροφητικά. Ταυτόχρονα, η ενισχυμένη τους ικανότητα οξείδωσης για As(III) σε As(V) μπορεί να μειώσει σημαντικά την τοξικότητα του As(III) στο περιβάλλον. Αυτοί οι ληφθέντες μπιρνεσίτες έχουν μεγάλες πιθανές εφαρμογές στην αποκατάσταση εδάφους και νερού που έχουν μολυνθεί από βαρέα μέταλλα. Ευχαριστίες Οι συγγραφείς ευχαριστούν θερμά το Εθνικό Ίδρυμα Φυσικών Επιστημών της Κίνας (Αριθμοί επιχορήγησης: 40830527, 41171375) και τα Ταμεία Βασικής Έρευνας για τα Κεντρικά Πανεπιστήμια (Αριθμός Προγράμματος: 2011PY015) για την οικονομική υποστήριξη. Οι συγγραφείς αναγνωρίζουν επίσης τον βοηθό ερευνητή Homer Genuino στο Πανεπιστήμιο του Κονέκτικατ για τη βελτίωση της αγγλικής γραφής στην εργασία. Παράρτημα A. Συμπληρωματικά δεδομένα Συμπληρωματικά δεδομένα που σχετίζονται με αυτό το άρθρο βρίσκονται στην ηλεκτρονική έκδοση στη διεύθυνση doi:10.1016/j.jhazmat.2011.09.027. Αναφορές [1] Υ.Ν. Vodyanitskii, Minerology and geochemistry of manganese: a review of publications, Eurasian Soil Sci. 42 (2009) 1170–1178. [2] F. Liu, C. Colombo, P. Adamo, J.Z. He, A. Violante, Ιχνοστοιχεία σε οζίδια μαγγανίου-σιδηρού από μια κινεζική Alfisol, Soil Sci. Soc. Είμαι. J. 66 (2002) 661–670. [3] R.P. Han, W.H. Ζου, Ζ.Π. Zhang, J. Shi, J.J. Yang, Αφαίρεση χαλκού(ΙΙ) και μολύβδου(ΙΙ) από υδατικό διάλυμα με επικαλυμμένη άμμο με οξείδιο μαγγανίου. Ι. Χαρακτηρισμός και κινητική μελέτη, J. Hazard. Μητήρ. B137 (2006) 384–395. [4] D. Mohan, C.U. Pittman Jr., Αφαίρεση αρσενικού από νερό/λύματα με χρήση προσροφητικών-Μια κριτική ανασκόπηση, J. Hazard. Μητήρ. 142 (2007) 1–53. [5] X.H. Feng, L.M. Zhai, W.F. Tan, F. Liu, J.Z. He, Προσρόφηση και οξειδοαναγωγικές αντιδράσεις βαρέων μετάλλων σε συνθετικά ορυκτά οξειδίου του Mn, Environ. Ρύπανση. 147 (2007) 366–373. [6] R.N. Dai, J. Liu, C.Y. Yu, R. Sun, Y.Q. Lan, J.D. Mao, Συγκριτική μελέτη οξείδωσης Cr(III) σε υδατικά ιόντα, σύμπλοκα ιόντα και αδιάλυτες ενώσεις από ορυκτό που φέρει μαγγάνιο (birnessite), Chemosphere 76 (2009) 536-541. [7] Y.T Meng, Y.M. Zheng, L.M. Zhang, J.Z. He, Biogenic Mn oxides for αποτελεσματική προσρόφηση Cd από υδάτινο περιβάλλον, Environ. Ρύπανση. 157 (2009) 2577–2583. [8] X.J. Li, C.S. Liu, F.B. Li, Υ.Τ. Li, L.J. Zhang, C.P. Liu, Y.Z. Zhou, The oxidative transformation of sodium arsenite at the interface of ␦-MnO2 and water, J. Hazard. Μητήρ. 173 (2010) 675–681. [9] S.B. Lee, J.S. An, Y.J. Kim, K. Nam, Binding force-associated toxicity reduce by birnessite and hydroxyapatite in Pb and Cd contaminated sediments, J. Hazard. Μητήρ. 186 (2011) 2117–2122. [10] R.G. Burns, V.M. Burns, Η ορυκτολογία και η κρυσταλλοχημεία των οζιδίων μαγγανίου βαθέων υδάτων - ένας πολυμεταλλικός πόρος του εικοστού πρώτου αιώνα, Philos. Μεταφρ. R. Soc. London Ser. Α286 (1977) 283–301. [11] R.M. Taylor, R.M. Mckenzie, Η συσχέτιση ιχνοστοιχείων με ορυκτά μαγγανίου σε αυστραλιανά εδάφη, Αυστ. J. Soil Res. 4 (1966) 29–39. [12] R.M. Mckenzie, Η αντίδραση του κοβαλτίου με ορυκτά διοξειδίου του μαγγανίου, Aust. J. Soil Res. 8 (1970) 97–106. [13] R.M. Mckenzie, Η προσρόφηση μολύβδου και άλλων βαρέων μετάλλων σε οξείδια μαγγανίου και σιδήρου, Aust. J. Soil Res. 18 (1980) 61-73. [14] R.G. Εγκαύματα, Η πρόσληψη κοβαλτίου σε οζίδια σιδηρομαγγανίου, εδάφη και συνθετικά οξείδια μαγγανίου (IV), Geochim, Cosmochim. Acta 40 (1976) 95–102.

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[40] V.A. Drits, E.J. Silvester, A.I. Gorshkov, A. Manceau, Structure of Na-rich synthetic monoclinic birnessite and hexagonal birnessite: I. X-ray diffraction and selected area electron diffraction results, Am. Mineral. 82 (1997) 946-961. [41] B. Lanson, V.A. Drits, E.J. Silvester, A. Manceau, Structure of H-altered hexagonal birnessite and its formation mechanism from Na-rich monoclinic busserite at low pH, Am. Mineral. 85 (2000) 826-838. [42] S.M. Webb, G.J. Dick, J.R. Bargar, B.M. Tebo, Evidence for the presence of Mn(III) intermediates in the bacterial oxidation of Mn(II), Proc. Natl. Acad. Sci. USA 102 (2005) 5558-5563. [43] J.J. Morgan, Manganese in natural waters and the Earth's crust: its availability to organisms, in: A. Sigel, H. Sigel (eds.), Metal Ions in Biological Systems, Manganese and its Role in Biological Processes, Marcel Dekker, New York, 2000 [ 44] S.W. Knipe, J.R. Mycroft, A.R. Pratt, H.W. Nesbitt, G.M. Bancroft, X-ray photoelectron spectroscopic study of water adsorption on iron sulfide minerals, Geochim. Cosmochim. Acta 59 (1995) 1079-1090. [45] A.S. Özcan, Ö. Gök, A. Özcan, Adsorption of lead(II) ions on 8 hydroxyquinoline immobilized bentonite, J. Hazard. Mater. 161 (2009) 499-509. [46] S.M. Maliyekkal, K.P. Lisha, T. Pradeep, A novel hybrid material of cellulose and manganese oxide by in situ soft chemical synthesis and its application for Pb(II) removal from water, J. Hazard. Mater. 181 (2010) 986-995. [47] C.H. Giles, T.H. MacEwan, S.N. Nakhwa, D. Smith, Adsorption studies. Part XI. A classification system for solution adsorption isotherms and its application for diagnosing adsorption mechanisms and measuring the specific surface area of solids, J. Chem. Soc. 3 (1960) 3973-3993. [48] D.G. Kinniburgh, Adsorption isotherms for general purposes, Environ. Sci. Technol. 20 (1986) 895-904. [49] R.P. He, W.H. Zhou, H.K. Lee, Y.H. Li, J. Shi, Removal of copper(II) and lead(II) from aqueous solution in manganese oxide-coated fixed-bed zeolite columns, J. Hazard. Mater. B137 (2006) 934-942. [50] E. Eren, B. Afsin, Y. Onal, Removal of lead ions by acid activated bentonite coated with manganese oxide, J. Hazard. Mater. 161 (2009) 677-685. [51] K.D. Kwon, K. Refson, G. Sposito, Surface complexation of Pb(II) by hexagonal birnessite nanoparticles, Geochim. Cosmochim. Acta 74 (2010) 6731-6740. [52] W. Zhao, Q.Q. Wang, F. Liu, G.H. Qiu, W.F. Tan, X.H. Feng, Pb2+ sorption on birnessite under the influence of Zn2+ and Mn2+ pretreatments, J. Soil Sediment. 10 (2010) 870-878. [53] M. Villalobos, J. Bargar, G. Sposito, Mechanisms of Pb(II) sorption on biogenic manganese oxide, Environ. Sci. Technol. 39 (2005) 569-576. [54] M.G. Keiser, W.H. Van Riemsdijk, ECOSAT: Equilibrium Calculation of Speciation and Transport, User Manual Version 4.7, Wageningen Agricultural University, The Netherlands, 1999. [55] A. Amirbahman, D.B. Kent, G.P. Curtis, J.A. Davis, Kinetics of arsenic (III) sorption and abiotic oxidation by aquifer materials, Geochim. Cosmochim. Acta 70 (2006) 533-547. [56] R.M. Mckenzie, The effect of cobalt on the reactivity of manganese dioxide, Aust. J. Soil Res. 9 (1971) 55-58. [57] B.J. Lafferty, M. Ginder-Vogel, M.Q. Zhu, K.J.T. Livi, D.L. Sparks, oxidation of arsenite from weakly crystalline manganese oxide. 2. X-ray absorption spectroscopy and X-ray diffraction results, Environ. Sci. Technol. 44 (2010) 8467-8472. [58] C. Tournassat, L. Charlet, D. Bosbach, A. Manceau, Arsenic(III) oxidation by birnessite and precipitation of manganese(II) arsenate, Environ. Sci. Technol. 36 (2002) 493-500. [59] M.Q. Zhu, K.W. Paul, J.D. Kubicki, D.L. Sparks, A quantum chemical study of arsenic(III,V) adsorption on Mn oxides: implications for arsenic(III) oxidation, Environ. Sci. Technol. 43 (2009) 6655–6661. [60] J.G. Speight, Lange's Handbook of Chemistry, 16th edition, McGraw-Hill Press, New York, 2005. [61] B.A. Manning, S.E. Fendorf, B. Bostick, D.L. Suarez, Arsenic(III) oxidation and adsorption reactions on synthetic birnessite, Environ. Sci. Technol. 36 (2002) 976-981. [62] X.H. Feng, W.F. Tan, F. Liu, H.D. Ruan, J.Z. He, Oxidation of AsIII by various manganese oxide minerals in the absence and presence of goethite, Acta Geol. A sin. (Eng.) 80 (2006) 249-256.

Journal of Hazardous Materials 196 (2011) 327-334

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Adsorption Behavior of Some Aromatic Compounds on Hydrophobic Magnetite for Magnetic Separation Takahiro Sasaki ∗, Shunitz Tanaka Department of Environmental Science, Faculty of Environmental Science, Hokkaido University, Sapporo, Hokkaido, 060-0810, Japan

article

i n f o

Article history: Received March 4, 2011 Received in revised form September 8, 2011 Accepted September 9, 2011 Available online September 16, 2011 Keywords: adsorption behavior Aromatic compounds Hydrophobic magnetite Hydrophobic interaction ␲ electron interaction

a b s t r a c t In this study, hydrophobic magnetite coated with an alkyl chain or phenyl group on the surface was prepared and used as an adsorbent to investigate the adsorption behavior of aromatic compounds with different log Pow values (phenol 1.46, benzonitrile 1.56, nitrobenzene ), benzene 2.13, toluene 2.73, chlorobenzene 2.84 and o-dichlorobenzene 3.38) on hydrophobic magnetite. Hydrophobic magnetites were modified with stearic acid and phenyltrimethoxysilane, and the amounts of modification were 9.84 × 10-3 and 4.17 × 10-2 mmol/g, respectively. Aromatic compounds used in this study are divided into 3 groups according to log Pow: 1 < log Pow < 2, 2 < log Pow < 3 and 3 < log Pow. The amounts of adsorption of each group on magnetite at an initial concentration of 100 ppm were 3.62 × 10-3 (nitrobenzene), 1.92 × 10-2 (phenol), 1.13 × 10-1 (chlorobenzene), 2.42 × 10-1 (benzene) or 3.10 × 10-1 mmol/g (dichlorobenzene). This indicates that the adsorption behavior depends on the hydrophobicity strength of the aromatic compounds. The adsorption mechanism for 2 < log Pow < 3 and 3 < log Pow is hydrophobic interaction, and that for 1 < log Pow < 2 is ␲-electron interaction. The quantitative relationship between the amount of adsorbed compounds and modified functional groups and suitability for isothermal adsorption models suggest that in most cases this adsorption could form multilayer adsorption. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Global attention to water pollution by harmful chemicals has increased in recent decades. Many types of pollutants have been detected in aquatic environments such as rivers, lakes and oceans. These pollutants originate from industrial and domestic sewage, sometimes from accidental spills. Various monocyclic and polycyclic aromatic compounds have been found in the aquatic environment. These aromatic compounds must be removed before the water is discarded or consumed. Currently, two types of technology are available for the treatment of harmful organic compounds in wastewater. The first is decomposition technology, where hazardous organic compounds are converted into more environmentally friendly compounds. Technologies such as chemical oxidation [1], electrolysis [2], photooxidation [3] and ozonization [4] are included in this category. The second type is separation technology, where harmful organic pollutants are separated from wastewater using different methods.

∗ Corresponding author at: N10W5 kita-ku, Sapporo, Hokkaido 060-0810, Japan. Tel.: +81 11 706 2219, fax: +81 11 706 2219. E-mail addresses:[email protected](T. Sasaki),[email protected](S. Tanaka). 0304-3894/$ – see cover page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.033

Technologies such as membrane separation [5], adsorption [6] and solvent extraction [7] belong to the second category. Among these types of technologies, adsorption is one of the simplest and most effective processes. Adsorption is a fast, economical and widely applicable technique. The use of adsorption is applicable to various pollutants such as organic compounds and heavy metals by selecting the type of adsorbent and adsorption conditions. In addition, wastewater treatment methods using low-cost adsorbents, such as by-products or waste materials, have recently been published [8,9]. Gupta et al reviewed the details of treatment methods for various pollutants in water using inexpensive adsorbents [10,11]. Activated carbon is the most widely used adsorbent in various cases due to its high capacity and wide range of adsorbents. However, there are certain limitations, especially in regeneration [12]. There is poor mechanical stiffness and low selectivity when activated carbon is used for actual environmental pollution. In addition, it is difficult to collect activated carbon dust that is widely dispersed in the environment. If not collected, the adsorbent used to adsorb harmful pollution can become a secondary source of pollution. Magnetic separation has recently been applied in various fields such as analytical biochemistry [13], medical science [14] and biotechnology [15]. From an environmental point of view, magnetic separation offers advantages due to its easy recovery

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1 < log Pow < 2 OH

1,46 a)

1,56 a)

Phenol

Benzonitril

NO2

1,62 d)

1,86 a)

Nitrocyclohexane

Nitrobenzene

2 < log Pow < 3 Cl

CH3

2.13 a)

2,73 b)

Benzene

Toluene

2.84 a) Chlorobenzene

3 < log Pow Cl Cl

3,38 c)

o-dichlorobenzene Fig. 1. Structures and log POW of aromatic compounds and nitrocyclohexane. a) log Pow values given in Ref [25] b) log Pow values given in Ref [26] c) log Pow values given in Ref [27] d) log Pow values given in Ref. [28].

adsorbent without filtration or centrifugation. Several studies have reported magnetic separation using modified magnetite (Fe3O4) as an environmentally friendly approach for the removal of heavy metal ions [16,17] and organic pollutants [18,19]. The removal of harmful organic compounds by adsorbents in most cases depends on the hydrophobic interaction between the adsorbent and its target compounds. This hydrophobic interaction has been used not only for the removal of harmful substances from adsorbents, but also for preconcentration of analytes using solvents [7] and extraction of solids [20]. A hydrophobic adsorbent can be prepared by using a hydrophobic agent on the surface of materials such as magnetite and silicate beads [18,21]. One of the most popular hydrophobization techniques involves the use of silane coupling agents. Silane coupling agents are known as surface modifiers that can add organic properties to the surface of an inorganic material [22,23]. Therefore, the silane coupling agent is important in the formation of the inorganic-organic hybrid material. Another technique is to use an ionic surfactant. The ionic group of the surfactant will be directed toward the surface of mineral oxides such as aluminum, silicon, titanium dioxide, and iron oxide, and then the alkyl chain, the hydrophobic group of the surfactant, will be directed outward. As a result, the surface of the material becomes hydrophobic [24]. The strength of the hydrophobic interaction depends on the degree of hydrophobicity of both the adsorbent and the adsorbate. The hydrophobicity of an organic compound can vary depending on the structure and functional group of the compound. The octanol-water partition coefficient (Pow) is a well-known indicator of the hydrophobicity of an organic compound. A compound with higher hydrophobicity will have a higher Pow. Therefore, the hydrophobicity, or Pow, of an organic compound is very important in predicting the adsorption behavior of certain organic compounds in water. The aim of this research was to clarify the adsorption behavior of organic compounds on hydrophobic magnetite and to evaluate the potential of magnetic separation for the removal of organic compounds dissolved in water. Organic compounds

with low Pow s, or relatively weak hydrophobicity, were used in this study. These organic compounds were selected based on their log Pow value. The selected compounds are divided into 3 groups according to log Pow: 1 < log Pow < 2, 2 < log Pow < 3 and 3 < log Pow. The Pow of phenol, benzonitrile and nitrobenzene decreased in group 1 < log Pow < 2. Those of benzene, toluene and chlorobenzene were in group 2 < log Pow < 3. o-dichlorobenzene was in group 3 < log Pow [25 -29 ] . The individual log Pows values of the aromatic compounds used in the adsorption experiments are summarized in the figure. 1. Hydrophobic magnetite is prepared by hydrophobizing the surface of the magnetite particle. Stearic acid and phenyltrimethoxysilane were used to hydrophobize the magnetite surface. By using two different types of hydrophobic magnetite, the difference in the adsorption behavior of different aromatic compounds was investigated. 2. Experiments 2.1. Magnetite materials whose average size was 0.3 µm (data provided by Kishida Chemical.) were purchased from Kishida Chemical Co., Ltd. (Osaka, Japan) and was used as an adsorbent carrier. The modifying reagents, stearic acid and phenyltrimethoxysilane were purchased from MP biomedicals Japan K.K. (Tokyo, Japan) and Tokyo Chemical Industry Co., Ltd. (Tokyo, Japan). The organic compounds used as adsorbents, phenol, nitrobenzene, and benzene were purchased from Wako Pure Chemical Industries, Ltd. (Osaka, Japan). Toluene, chlorobenzene, and o-dichlorobenzene were purchased from Nacalai Tesque, Inc. (Kyoto, Japan). Benzonitrile was purchased from Kanto Chemical Co., Inc. (Tokyo, Japan). Nitrocyclohexane was purchased from Tokyo Chemical Industry Co., Ltd. In fatty acid analysis, boron trifluoride solution was used as an esterifying agent, anhydrous sodium sulfate was used as a dehydrating agent, and methylene chloride was used as a solvent, all of which were purchased from Wako Pure Chemical Industries, Ltd. Reagents used for Si analysis (hydrochloric acid, nitric acid,

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θειικό οξύ, υδροφθορικό οξύ, βορικό οξύ, τετραένυδρο επταμολυβδαινικό εξααμμώνιο και χλωριούχο νάτριο) αγοράστηκαν από την Wako Pure Chemical Industries, Ltd. Όλα τα αντιδραστήρια ήταν αναλυτικής ποιότητας και χρησιμοποιήθηκαν χωρίς περαιτέρω καθαρισμό. 2.2. Παρασκευή υδρόφοβου μαγνητίτη Παρασκευάστηκαν δύο είδη υδρόφοβων προσροφητών: μαγνητίτης τροποποιημένος με στεατικό οξύ (SA-mag) και τροποποιημένος μαγνητίτης φαινυλομάδας (Ph-mag) [29,30]. Το SA-mag παρασκευάστηκε με τον ακόλουθο τρόπο. Προστέθηκε 1,0 g μαγνητίτη σε 50 ml μεθανόλης για διανομή. Στη συνέχεια προστέθηκαν 10 mg στεατικού οξέος στο εναιώρημα με ανάδευση για να στεγνώσει η μεθανόλη. Το υπόλειμμα πλύθηκε δύο φορές με μεθανόλη και ξηράνθηκε σε φούρνο στους 50°C. Το Ph-mag παρασκευάστηκε με τον ακόλουθο τρόπο. Προστέθηκε 1,0 g μαγνητίτη σε 40 ml αιθανόλης και κατανεμήθηκε με ανάδευση. Στη συνέχεια, 0,1 ml φαινυλοτριμεθοξυσιλανίου προστέθηκε στο εναιώρημα. Αφού το φαινυλτριμεθοξυσιλάνιο διαλύθηκε επαρκώς στο διαλύτη, προστέθηκαν στο εναιώρημα 0,058 ml Η2Ο και 0,025 ml 1 Μ HCl. Το εναιώρημα στη συνέχεια θερμάνθηκε στους 50 ◦ C με συνεχή ανάδευση μέχρι να στεγνώσει ο διαλύτης. Ο λαμβανόμενος μαγνητίτης θερμάνθηκε σε κλίβανο σιγαστήρα στους 120 ◦ C για 1 ώρα. Μετά τη θέρμανση, ο τροποποιημένος μαγνητίτης πλύθηκε δύο φορές με αιθανόλη και ξηράνθηκε σε φούρνο στους 50 ◦ C. 2.3. Χαρακτηρισμός υδρόφοβου μαγνητίτη 2.3.1. Επιφάνειες υδρόφοβου μαγνητίτη Τα εμβαδά επιφάνειας του SA-mag και του Ph-mag προσδιορίστηκαν χρησιμοποιώντας ανάλυση N2 BET (AutoSorb6 YUASA, Ιαπωνία). Τα δείγματα υποβλήθηκαν σε προεπεξεργασία με απαέρωση στους 80 ◦ C για 6 ώρες. 2.3.2. Ποσότητα τροποποιημένου στεατικού οξέος σε μαγνητίτη Για εκρόφηση τροποποιημένου στεατικού οξέος από SA-mag, 30 mg SA-mag πλύθηκαν με 20 ml αιθανόλης για 1 ώρα με υπερήχηση. Η διαδικασία πλύσης επαναλήφθηκε 3 φορές. Και στη συνέχεια, το συλλεγμένο διάλυμα αιθανόλης που περιελάμβανε στεατικό οξύ εξατμίστηκε υπό κενό. Το ληφθέν δείγμα διαλύθηκε σε 3 ml μεθανόλης και 1 ml μεθανολικού διαλύματος που περιείχε λαυρικό οξύ (1 mg/l) ως υποκατάστατο πρότυπο. Το μίγμα μεταφέρθηκε σε δοκιμαστικό σωλήνα και προστέθηκαν 2 ml μεθανολικού διαλύματος 14% BF3. Το δείγμα τοποθετήθηκε σε υδατόλουτρο στους 80 ◦ C για 3 λεπτά και στη συνέχεια προστέθηκε 1 ml νερού για να σταματήσει η αντίδραση. Οι εστέρες λιπαρών οξέων εκχυλίστηκαν δύο φορές με 1 ml μεθυλενοχλωριδίου κάθε φορά. Η οργανική φάση που συλλέχθηκε αφυδατώθηκε με άνυδρο θειικό νάτριο και τοποθετήθηκε σε ογκομετρική φιάλη 10 ml μετά από διήθηση. Τα προσαρμοσμένα 10 ml διαλύματος δείγματος μετρήθηκαν χρησιμοποιώντας GC-17A (Shimadzu, Japan) εξοπλισμένο με GCMS-QP5050A (Shimadzu, Japan). Χρησιμοποιήθηκε στήλη DB-5 ms (30 m x 0,25 mm x 0,25 ␮m) (Agilent, USA). 1 ␮l διαλύματος δείγματος εγχύθηκε. Οι ενέσεις πραγματοποιήθηκαν σε λειτουργία χωρίς διαχωρισμό. Το φέρον αέριο ήταν ήλιο (Air Water, Ιαπωνία) με σταθερή ροή 1,5 ml/min. Η θυρίδα έγχυσης θερμάνθηκε στους 200 ◦ C. Η θερμοκρασία του φούρνου ρυθμίστηκε στους 75 ◦ C για 2 λεπτά, στη συνέχεια αυξήθηκε 30 ◦ C/min στους 270 ◦ C και η τελική θερμοκρασία διατηρήθηκε για 1 λεπτό. Η θερμοκρασία στον ανιχνευτή ήταν 280 ◦ C. Όλα τα φάσματα μάζας αποκτήθηκαν στον τρόπο κρούσης ηλεκτρονίων (EI) ως πηγή ιονισμού με ένα τετραπολικό φίλτρο μάζας. Η ανάλυση διεξήχθη σε λειτουργία SIM και τα επιλεγμένα ιόντα της ένωσης ήταν m/z 55, 74 και 87. Οι συγκεντρώσεις των λιπαρών οξέων υπολογίστηκαν χρησιμοποιώντας τη μέθοδο εσωτερικού προτύπου [31]. 2.3.3. Ποσότητα τροποποιημένης φαινυλομάδας στον μαγνητίτη Η ποσότητα τροποποιημένου φαινυλοτριμεθοξυσιλανίου στην επιφάνεια του μαγνητίτη προσδιορίστηκε από τη μέτρηση του διοξειδίου του πυριτίου ως συστατικού αποσύνθεσης του Ph-mag με τον ακόλουθο τρόπο [32]. Αρχικά, 1 g Ph-mag αποσυντέθηκε χρησιμοποιώντας 40 ml του μικτού οξέος που περιείχε 6 Μ υδροχλωρικό οξύ και 6 Μ νιτρικό οξύ

329

cover the watch glass while heating to 80 ◦ C until the black precipitate disappears. The resulting precipitated silicate was filtered with a membrane filter and washed several times with water. Second, the filter paper carrying the precipitate was moved to a PTFE beaker, and then 5 mL of 4% (w/v) NaCl solution and 3 mL of 60% hydrofluoric acid were added to the beaker. The mixture was heated to dryness in a water bath. After dissolving the obtained residue with 15 ml of water, 10 ml of saturated boric acid solution was added to the mixture, and then it was heated almost to the boiling point (toluene > chlorobenzene. This result did not agree with the series of log Pow. That is, the order of adsorption capacity is not always depended on the log POW of the compounds of this group, the strength of interaction and the size of the molecule forming multilayer benzene was adsorbed more because there was no substituent group and it has the smallest molecular size [33] On the other hand, toluene and chlorobenzene are larger than benzene because of the substituent group. , the substituent group of toluene, donates electrons to the benzene ring because

A 0,25 qe (mmol/g)

qe (mmol/g)

A 0,40

331

Benzen Toluen Klorbenzen

0,20 0,15 0,10 0,05 0

Ce (ppm)

0,40

B 0,25

Ph-mag

0,30

qe (mmol/g)

SA-mag

0,20 0,10 0

0,20 0,15 0,10 0,05

Ce (ppm) Fig. 6. Adsorption amounts of o-dichlorobenzene on SA-mag and Ph-mag at different initial concentrations (10-100 ppm).

Ce (ppm)

Fig. 7. Amounts of adsorption of aromatic compounds at 2 < log Pow < 3 on (A) SA-mag and (B) Ph-mag at different initial concentrations (10-100 ppm).

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A functional group is an electron-donating group. In contrast, chlorine, the substituent group on chlorobenzene, is an electron-withdrawing group that withdraws an electron from the benzene ring. The strength of the ␲ electron interaction depends on the abundance of ␲ electrons in the benzene ring, i.e. the interaction works well in the electron-rich ␲ state. ␲ electron interaction between aromatic compounds appears to be important for the formation of multilayers in these cases. Therefore, toluene was adsorbed more than chlorobenzene. 3.3.3. 1 < log Pow < 2 Phenol, benzonitrile, nitrobenzene and nitrocyclohexane had the lowest hydrophobicity (1 < log Pow < 2) in this study. Fig. Fig. 8 shows the results of these adsorption experiments. The amounts of adsorption of these compounds in this group were small compared to other groups. This happened because these compounds have strong polar groups such as hydroxyl, cyano and nitro groups and can hydrate with water molecules around their polar groups. Therefore, the hydrophobic interaction between these compounds and the modified magnetite becomes weak. Similarly, the interaction between their junctions also becomes weak. According to Fig. 8, phenol, benzonitrile and nitrobenzene were more selectively adsorbed on Ph-mag than on SA-mag. These results show that the functional group on the magnetite surface significantly affects the adsorption behavior and ␲ electron interactions between the compounds and the phenyl group in Ph-mag are stronger than the hydrophobic interaction in these cases. When nitrocyclohexane, which does not have an aromatic ring, was weakly adsorbed on Ph-mag, it suggested that the adsorption mechanism for this group is based on ␲ electron interaction. The order of the amount of adsorption of these aromatic compounds was as follows: phenol > benzonitrile > nitrobenzene. These results did not agree with the order of log Pow (phenol 1.46, benzonitrile 1.56, nitrobenzene 1.86). Phenol has a hydroxyl group, an electron donating group and ␲ electron interaction between phenols

qe (mmol/g)

0,03

Fenol Benzonitril Nitrobenzen Nitrocikloheksan

0,02

and among them was a stronger Ph-mag. However, the substituent groups of benzonitrile and nitrobenzene are electron-withdrawing groups, which makes their aromatic rings relatively electron-poor. The electron-withdrawing property of benzonitrile may be weaker than that of nitrobenzene because the dipole moment of benzonitrile, 4.05, is smaller than that of nitrobenzene, 4.22–4.91 [34,35]. The aromatic ring of benzonitrile is more electron rich than the ring of nitrobenzene. The ␲ electron interaction between the compound and the phenyl group on the magnetite surface was dominant in this adsorption mechanism, and therefore the amount of adsorption of benzonitrile on Ph-mag was greater than that of nitrobenzene. According to fig. In Figure 8A, phenol was adsorbed on SA mag, and the amount of adsorbed phenol was close to the modified amount of stearic acid in magnetite. However, this does not mean that the adsorption is a one-to-one ratio between phenol and stearic acid as monolayer adsorption due to the shape of the adsorption isotherm. The characteristic of isothermal adsorption is that the amount of adsorption immediately increases with a high concentration of adsorbent during multilayer adsorption. In addition, the dipole moment for phenol, 1.53 [36], was the smallest, and the hydration for phenol was the weakest. Therefore, a small amount of phenol could be adsorbed on SA mag by hydrophobic interaction. However, the adsorption behavior of this group was not clear in the initial concentration range (10-100 ppm). Fig. Figure 8 shows that the adsorption amounts of the compounds are smaller compared to the amounts of modified functional groups, especially in the case of the phenyl group with an initial concentration of 100 ppm. Therefore, the adsorption experiment for this group should be performed with a higher initial concentration. 3.4. Isothermal Adsorption Models This section describes the investigation of adsorption isotherms for these aromatic compounds on two adsorbents (in Figure 5) using adsorption isotherm models. The model of multilayer adsorption in the gas phase is known as the BET model. Many researchers have reported the use of a liquid-phase multilayer adsorption model based on the BET model, but this model has not yet been established [37]. Therefore, in this study, the adsorption behavior was investigated by fitting these adsorption data to the Langmuir and Freundlich models. The Langmuir and Freundlich models are as follows: 1 1 1 = + qe KL qm Ce qm

0,01

1/n

qe = Kf Ce

0,00

0,03

qe (mmol/g)

0,02

60 Ce (ppm)

100

120

60 80 Ce (ppm)

100

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120

0,01

0,00 0

Fig. 8. Amounts of adsorption of aromatic compounds and nitrocyclohexane at 1 < log Pow < 2 on (A) SA-mag and (B) Ph-mag at different initial concentrations (10-100 ppm).

(1) (2)

where qe is the equilibrium adsorption capacity, Ce represents the equilibrium solute concentration, qm (mg/g) is the maximum adsorption capacity, KL (L/g) is the binding constant, and Kf and n are the Freundlich constants to be determined. Table 1 shows the fitting results for the Langmuir and Freundlich models. When using the Langmuir model, most correlation coefficients were not improved. The Langmuir model is well known as a monolayer adsorption model, and most of the adsorption behavior appears to be non-monolayer adsorption due to low correlation coefficients. On the other hand, when using the Freundlich model, most of the correlation coefficients are better than those of the Langmuir model. However, most of the correlation coefficients for the Freundlich model were not as good. The Freundlich model does not provide insight into the adsorption behavior due to the empirical formula. However, the Freundlich model is known to best fit adsorption on a porous material such as activated carbon. Therefore, these adsorption behaviors appear not to be porous adsorption. This result is also consistent with Figure 4. These fitting results for the models were not conclusive as to whether

T. Sasaki, S. Tanaka / Journal of Hazardous Materials 196 (2011.) 327–334

333

Table 1. Isothermal parameters for aromatic compounds in SA-mag and Ph-mag. Adsorbent

Adsorber

Friendly model

Langmuirov model

qm (mmol/L)

KL (L/mg)

SA-mag

fenol benzonitril nitrobenzen benzen toluen klorbenzen o-diklorbenzen

3,46e−5 2,70e−4 – 3,83 1,76 2,80 16,4

0,453 0,647 – 2,52 1,80 1,45 2,97

0,807 0,988 – 0,984 0,962 0,930 0,883

1,17e−3 9,52e−4 – 0,234 0,203 0,0690 0,248

0.0101 9.86e−3 – 0.135 0.0543 0.933 1.90

0,213 0,991 – 0,960 0,911 0,423 0,979

Ph-mag

fenol benzonitril nitrobenzen benzen toluen klorbenzen o-diklorbenzen

0,528 0,0383 0,115 3,78 1,53 1,56 16,1

3,63 1,26 1,83 3,07 1,70 0,885 2,82

0,932 0,978 0,995 0,951 0,935 0,986 0,853

0,0189 0,0185 0,0134 0,135 0,199 0,0907 0,292

0,135 0,0151 0,0310 0,508 0,0534 0,224 1,19

0,819 0,937 0,908 0,812 0,927 0,724 0,944

0,05 Ph-mag

qe (mmol/g)

0,04

SA-mag

0,03 0,02 0,01 0

Inch. salt (w/v%) Fig. 9. Effect of salt on the adsorption of nitrobenzene on SA-mag and Ph-mag.

The adsorption behavior at 1 < log Pow < 2 was monolayer or multilayer adsorption. 3.5. Adsorption of nitrobenzene at different salt concentrations Nitrobenzene adsorption isotherms for different salt concentrations on SA-mag and Ph-mag are shown in fig. 9. This experiment was performed to confirm that water hydration inhibits the adsorption of aromatic compounds at 1 < log Pow < 2 on hydrophobic magnetite. The dehydration effects of adding salt were investigated. The addition of salt deprives the hydrated adsorbent of water molecules. If the adsorption amount of nitrobenzene increases when the adsorption experiment is performed under dehydration conditions, hydration will be an inhibiting factor in its adsorption on hydrophobic magnetite. As shown in fig. 9, the adsorption amounts of nitrobenzene on SA-mag and Ph-mag increase with increasing salt concentration. Moreover, nitrobenzene was almost not adsorbed on SA mag without the addition of salt, but it was adsorbed on SA mag in the presence of salt. The adsorption of nitrobenzene on SA mag was by hydrophobic interaction, which seems to indicate that the hydrophobicity of nitrobenzene is increased by dehydration. This result indicates that hydration is one of the inhibitors of nitrobenzene adsorption on hydrophobic magnetite. 4. Conclusion Aromatic compounds with different log Pow were tested for adsorption on alkyl chains and phenyl groups coated with hydrophobic magnetite. The adsorption behavior of each compound is divided into 3 groups according to log Pow: 1 < log Pow < 2, 2 < log Pow < 3 and 3 < log Pow. Adsorption sizes in general

increased as the log Pow of each group increased. However, the adsorption amounts for the compounds in each group did not depend on the log Pow values. In groups 2 < log Pow < 3 and 3 < log Pow, the adsorption mechanism was mainly hydrophobic interaction between aromatic compounds and the hydrophobic surface of magnetite. Adsorption behavior did not depend on the difference of modified functional groups on the magnetite surface. The adsorption behavior of hydrophobic magnetite on these groups appears to create a multilayer layer on hydrophobic magnetite due to the quantitative ratio between the amount of adsorbed aromatic compound and the amount of modified functional groups on magnetite, and there was a poor fit for the adsorption isotherm models. However, the adsorption behavior for group 1 < log Pow < 2 was sensitive to the modified functional group on the hydrophobic magnetite. The main adsorption mechanism for this group was the ␲-electron interaction between the compounds and the phenyl group on the Ph-mag rather than the hydrophobic interaction. The adsorption behavior of this group could not be proven either for single-layer or multilayer adsorption under these adsorption conditions. The results of adsorption experiments under dehydration conditions showed that the inhibiting factor for nitrobenzene adsorption is the hydration of water molecules. The simple system in this study, which used non-porous magnetite, allowed elucidation of the factors that determine the adsorption behavior. The results were therefore important in terms of selecting the optimal surface modifier for solid phase adsorption or extraction as well as magnetite separation, which could be valuable in the design of a new high efficiency adsorbent. Acknowledgments. The N2 BET analysis in this paper was performed with Autsorb6 at the OPEN FACILITY, Hokkaido University Sousei Hall. Bibliography [1] E. Ferrarese, G. Andreottola, I.A. Oprea, Remediation of PAH-contaminated sediments by chemical oxidation, J. Hazard. Mater. 152 (2008) 128-139. [2] Z.M. Shen, D. Wu, J. Yang, T. Yuan, W.H. Wang, J.P. Jia, Methods for improving the performance of electrochemical treatment of dye wastewater, J. Hazard. Mater. 131 (2006) 90-97. [3] X. Shen, L. Zhu, G. Liu, H. Yu, H. Tang, Enhanced photocatalytic degradation and selective removal of nitrophenol using surface molecularly imprinted titanium, Environ. Sci. Techol. 42 (2008) 1687-1692. [4] F.J. Beltrán, G. Ovejero, J.M. Encinar, J. Rivas, Oxidation of polynuclear aromatic hydrocarbons in water. 1. Ozonization, Ind. Meadow. Chem. Crisp. 34 (1995) 1596-1606. [5] UK Ghosh, N.C. Pradhan, B. Adhikari, Separation of water and o-chlorophenol by evaporation using HTPB-based polyurethane membranes and application of the modified Maxwell-Stefan equation, J. Membr. Sci. 272 (2006) 93–102.

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Journal of Hazardous Materials 196 (2011) 335-341

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Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Zinc stimulates the release of chemokines and inflammatory cytokines from human promonocytes Tsui-Chun Tsou a,∗, How-Ran Chao b, Szu-Ching Yeh a, Feng-Yuan Tsai a, Ho-Jane Lin a a b

Department of Environmental Health and Occupational Medicine, National Health Research Institute, Zhunan, Miaoli 350, Taiwan, Department of Environmental Science and Engineering, National Pingtung University of Science and Technology, Neipu, Pingtung 912, Taiwan

article

i n f o

Article history: Received May 18, 2011 Received in revised form September 8, 2011 Accepted September 9, 2011 Available online September 19, 2011 Keywords: zinc promonocytes Chemokines Inflammatory cytokines

a b s t r a c t Our previous studies showed that zinc oxide (ZnO) particles induce intercellular adhesion molecule-1 (ICAM-1) protein expression in vascular endothelial cells via NF-␬B, and that zinc ions dissolved in ZnO particles may play a major role in the process processing. This study aimed to determine whether zinc ions can induce inflammatory reactions in the human promonocytic leukemia cell line HL-CZ. Adapted medium from zinc-treated HL-CZ cells induced ICAM-1 protein expression in human umbilical vein endothelial cells (HUVECs). Zinc treatment induced the release of chemokines and inflammatory cytokines from HL-CZ cells. Inhibition of NF␬B activity by overexpression of I␬B␣ in HL-CZ cells did not prevent conditioned medium-induced ICAM-1 protein expression in HUVEC cells. Zinc treatment induced the activation of several immune response-related transcription factors in HL-CZ cells. These results clearly show that zinc ions induce the release of chemokines and inflammatory cytokines from human promonocytes, accompanied by the activation of several immune response-related transcription factors. Our in vitro evidence for zinc-induced vascular inflammatory responses provides a critical link between zinc exposure and the pathogenesis of these inflammatory vascular diseases. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Epidemiological studies show that exposure to fine particles (PM) in polluted air is associated with systemic inflammatory markers [1] as well as the occurrence of cardiovascular morbidity and mortality [2-4]. However, the mechanisms behind this association remain largely unknown. When inhaled into the respiratory tract, small particles tend to land in the deepest part of the lungs. Ultrafine particles [5-7] and, of particular importance for this study, their dissolved chemicals, such as metal ions, can penetrate deep into the lung and cross the lung epithelial barrier into the bloodstream, directly exposing vascular cells such as monocytes and endothelial cells. cell, in pollutants. Analysis of the level of zinc in PM10 particles in the air showed that 82-93% of zinc is found in small PM10 particles [8]. Studies of mothers who were exposed to cigarette smoking or air pollution showed that both smoking and air pollution contributed to an increase in zinc levels in the placenta [9]. A previous study of airborne zinc levels and health care utilization for asthma found an association between elevated airborne zinc and increased

∗ Corresponding author. Phone: +886 37 246 166x36511; fax: +886 37 587 406. E-mail address:[email protected](T.-C. Tsou). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.035

morbidity of pediatric asthma [10]. ZnO particles induce cytotoxicity and apoptosis in mammalian cells [11,12], and dissolved zinc ions seem to play a key role in the toxic effect of ZnO particles [13]. In our previous studies, ZnO particles induced ICAM-1 expression in vascular endothelial cells through an NF␬B-dependent pathway [14], and zinc ions alone were sufficient to induce similar levels of ICAM-1 expression as ZnO particles, suggesting that dissolved zinc ions may play a major role in the inflammatory effect of ZnO particles on vascular endothelial cells [15]. These studies show that the composition of metal particles or their dissolved metal ions can determine the ability of metal oxide nanoparticles to induce inflammation in vascular endothelial cells. Increasing evidence indicates that chronic obstructive pulmonary disease, asthma, and atherosclerosis are associated with systemic inflammatory cytokine changes. A variety of pathophysiological stimuli induce cytokine release, including modified LDL [16,17] , free radicals [18] , hemodynamic stress [19,20] , and hypertension [21] . Based on our previous findings in vascular endothelial cells using ZnO particles, this study aimed to determine whether zinc can induce inflammatory responses in other vascular cells. We found that zinc induces the release of chemokines and inflammatory cytokines from human promonocytes, possibly through the activation of several immune response-related transcription factors.

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2. Materials and methods 2.1. Materials Zn(CH3COO)2 (#370080250) were obtained from ACROS Organics (Geel, Belgium). The zinc preparation in this study was tested for endotoxin-free using the endotoxin inhibitor, polymyxin B, as previously described [22]. RayBio Human Cytokine Antibody Array 3 (#AAH-CYT-3) was purchased from RayBiotech, Inc. (Norcross, GA, USA). Rabbit polyclonal antibodies against ICAM-1 (sc-7891) and I␬B␣ (sc-847) and goat polyclonal antibodies against p-I␬B␣ (sc-7977) were purchased from Santa Cruz Biotechnology (Santa Cruz, CA, USA). ). Mouse monoclonal anti-actin antibody (MAB1501) was purchased from Chemicon Int. Inch. (Temecula, CA, USA). Endothelial cell growth supplement (ECGS) was obtained from Sigma-Aldrich (St. Louis, MO, USA). Dulbecco's phosphate buffered saline (D-PBS), M199 medium and RPMI 1640 medium were obtained from Life Technologies (Grand Island, NY, USA). Fetal bovine serum (FBS) was obtained from HyClone (Logan, UT, USA). Penicillin (10,000 units/ml)/streptomycin (10,000 μg/ml) solution was purchased from Invitrogen Corp. (Carlsbad, CA, USA). Gentamicin sulfate was obtained from Biological Industries (Kibbutz Beit Haemek, Israel). 2.2. Construction of recombinant adenoviruses Construction of recombinant AdEasy-GFP, AdEasy-I␬B␣ and AdV-NF␬B-Luc was previously described [14]. To construct AdV-AP-1-Luc, a 2360 bp DNA fragment containing seven copies of the AP-1 response element, the TATA box and the gunpowder luciferase gene in pAP-1-Luc (Stratagene, La Jolla, CA, USA) , was amplified by PCR using pfu DNA polymerase. The amplified DNA fragment was digested with KpnI and SalI. After separation by agarose gel electrophoresis, the purified KpnI/Sall-digested DNA fragment was cloned into the KpnI/Sall-digested pACCMV.pLpA vector [23]. The CMV promoter of the pACCMV.pLpA vector used here was removed. Recombinant adenovirus AdVAP-1-Luc was generated by homologous recombination between plasmid pJM17 [24] and vector pACCMV.pLpA in 293 human embryonic kidney cells. Construction of other recombinant adenoviruses (AdEasy-C/EBP-Luc, AdEasy-CRE-Luc, AdEasy-NFATLuc, AdEasy-SRE-Luc, and AdEasy-STAT-Luc) were generated using the AdEasyTM adenoviral vector system, La Jollagene, CA, USA) (see Supplementary material for details). Recombinant adenoviruses were purified and concentrated according to the manufacturer's instructions. General information about these immune-related transcription factor-linked luciferase reporter adenoviruses is summarized in Table 1. 2.3. Cells and therapies Human promonocytic leukemia cell line HL-CZ (BCRC-60043), originally established by Dr. Wu-Tse Liu (National Yang-Ming University, Taipei, Taiwan) [25], were purchased from the Bioresource Collection and Research Center (BCRC, Hsinchu, Taiwan) and routinely cultured in RPMI 1640 medium. HUVEC cells were obtained by digestion with vein collagenase umbilical cords [26] and routinely cultured in M199 medium as previously described [27]. HL-CZ cells (8.5 x 10 6 cells per 100 mm dish) were left untreated or treated with Zn(CH 3 COO) 2 as indicated for 6 h. After treatment, the conditioned medium was dialyzed against D-PBS with agitation at 4°C for 42 h, sterilized with a 0.45-␮m syringe filter, and then supplemented with M199 medium (with 20% FBS and 30 µSg/ml ECGS) in a ratio of 1/1 (v/v). HUVEC cells were treated with these conditioned medium/M199 mixtures for various time periods

Οπως αναφερεται. Μετά τις θεραπείες, συλλέχθηκαν προϊόντα λύσης κυττάρων για αναλύσεις ανοσοστύπωσης. Σε ορισμένες περιπτώσεις που απαιτούσαν μόλυνση από αδενοϊό, τα κύτταρα HL-CZ (3 × 106 κύτταρα ανά δίσκο 100 mm) μολύνθηκαν για πρώτη φορά με AdEasyGFP ή AdEasy-I␬B␣ σε πολλαπλότητα μόλυνσης (MOI) 50 pfu/κύτταρο για 24 ώρες . Τα μολυσμένα κύτταρα αντικαταστάθηκαν με φρέσκο μέσο RPMI 1640 και καλλιεργήθηκαν για άλλες 24 ώρες για ανάκτηση. Στη συνέχεια, τα κύτταρα ήταν έτοιμα για τις επεξεργασίες ψευδαργύρου όπως μόλις περιγράφηκε. 2.4. Ανάλυση ανοσοστυπώματος Μετά από θεραπείες, τα κύτταρα λύθηκαν σε παγωμένο ρυθμιστικό διάλυμα RIPA (50 mM Tris-HCl, pH 7,5, 5 mM EDTA, 1 mM EGTA, 1% Triton X-100, 0,25% δεοξυχολικό νάτριο) που περιείχε PMSF, (2 mM ), απρωτινίνη (2 ␮g/ml), λευπεπτίνη (2 ␮g/ml), NaF (2 mM), Na3VO4 (2 mM) και ␤-γλυκεροφωσφορικό (0,2 mM). Τα κυτταρολύματα υποβλήθηκαν σε ανάλυση SDS-PAGE και ανοσοστύπωμα, όπως περιγράφηκε προηγουμένως [28]. Τα στυπώματα ανιχνεύθηκαν με ένα πρωτεύον αντίσωμα έναντι του ICAM-1, φωσφόρου-Ι␬Β␣ (ρ-Ι␬Β␣), Ι␬Β␣, ή ακτίνης. Δευτερεύοντα αντισώματα συζευγμένα με HRP. Οι ζώνες πρωτεΐνης στη μεμβράνη οπτικοποιήθηκαν σε φιλμ ακτίνων Χ χρησιμοποιώντας Western Lightning Chemiluminescence Reagent Plus (PerkinElmer Life Sciences, Boston, MA, USA). Η ένταση της ζώνης πρωτεΐνης προσδιορίστηκε ποσοτικά με σάρωση πυκνομετρίας φιλμ ακτίνων Χ. 2.5. Ανάλυση κυτοκινών σε ρυθμισμένα μέσα από κύτταρα HL-CZ Μετά από επεξεργασίες με ψευδάργυρο των κυττάρων HL-CZ, τα ρυθμισμένα μέσα υποβλήθηκαν σε διαπίδυση έναντι D-PBS με ανάδευση στους 4 ◦ C για 42 ώρες και αποστειρώθηκαν με ένα φίλτρο σύριγγας 0,45 ␮m. Οι κυτοκίνες σε ρυθμισμένο μέσο αναλύθηκαν με το RayBio Human Cytokine Antibody Array 3 σύμφωνα με τις οδηγίες του κατασκευαστή (βλ. Συμπληρωματικό υλικό λεπτομερώς). Το κιτ παρέχει μια απλή μορφή πίνακα και μια εξαιρετικά ευαίσθητη προσέγγιση για την ταυτόχρονη ανίχνευση 42 επιπέδων έκφρασης κυτοκινών από ρυθμισμένα μέσα. 2.6. Δοκιμασία αναφοράς λουσιφεράσης που σχετίζεται με μεταγραφικό παράγοντα που σχετίζεται με την ανοσοαπόκριση Για να προσδιοριστεί η ενεργοποίηση αυτών των παραγόντων μεταγραφής που σχετίζονται με την ανοσοαπόκριση, κύτταρα HL-CZ (1,0 × 104 κύτταρα ανά φρεάτιο σε πλάκες 96 φρεατίων) μολύνθηκαν με έναν από τους ανασυνδυασμένους αδενοϊούς ( AdV-AP-1-Luc, AdV-NF␬B-Luc, AdEasy-SRE-Luc, AdEasy-NFATLuc, AdEasy-CRE-Luc, AdEasy-C/EBP-Luc και AdEasy-STAT-Luc) (Πίνακας 1 ) σε MOI 1 pfu/κύτταρο για 24 ώρες. Μετά τη μόλυνση με αδενοϊό, τα μολυσμένα κύτταρα αντικαταστάθηκαν με φρέσκο μέσο RPMI 1640 και καλλιεργήθηκαν για άλλες 24 ώρες για ανάκτηση. Στη συνέχεια, τα μολυσμένα κύτταρα αφέθηκαν χωρίς επεξεργασία ή υποβλήθηκαν σε επεξεργασία με 150 ␮M Zn(CH3COO)2 για 6 ώρες. Η δραστηριότητα της λουσιφεράσης κάθε δείγματος προσδιορίστηκε χρησιμοποιώντας το Σύστημα Δοκιμασίας Λουσιφεράσης (Promega, Madison, WI), σύμφωνα με τις οδηγίες του κατασκευαστή. 2.7. Στατιστικά Κάθε πείραμα πραγματοποιήθηκε ανεξάρτητα τουλάχιστον τρεις φορές. Η στατιστική ανάλυση εκφράστηκε χρησιμοποιώντας τη μέση ± τυπική απόκλιση (SD) από κάθε ανεξάρτητο πείραμα. Η επαγωγή της έκφρασης της πρωτεΐνης ICAM-1 σε κύτταρα HUVEC από ρυθμισμένο μέσο εξετάστηκε με δοκιμές t Student με 2000 δείγματα bootstrap. Χρησιμοποιήθηκαν δοκιμές t ενός δείγματος για τον προσδιορισμό των σημαντικών διαφορών στην επαγωγή χημειοκίνης και απελευθέρωσης φλεγμονώδους κυτοκίνης από κύτταρα HL-CZ μεταξύ των ομάδων που υποβλήθηκαν σε αγωγή με ψευδάργυρο και των ομάδων που δεν υποβλήθηκαν σε θεραπεία (τιμή δοκιμής = 1). Οι διαφορές θεωρήθηκαν στατιστικά σημαντικές όταν p < 0,05. Οι αναλύσεις πραγματοποιήθηκαν χρησιμοποιώντας το Statistical Package for Social Sciences (SPSS) έκδοση 12.0 (SPSS Inc., Chicago, IL, USA).

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Table 1 Transcription factor-mediated luciferase reporter adenoviruses associated with immunological response. Recombinant adenoviruses

Transcription factors

Response element sequences (RES) in one direction (5 → 3)

AdV-AP-1-Luc

AP-1

TGACTAATGACTAATGACTAATGACTAA TGACTAATGACTAATGACTAA

AdEasy-C/EBP-Luc

C/EBP

ATTGCGCAATATTGCGCAATATTGCGCAAT

AdEasy-CRE-Luc

CREBP

AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAGAGCCTGACGTCAGAG AGCCTGACGTCAGAG

AdEasy-NFAT-Luc

NFAT

ACGCCTTCTGTATGAAACAGTTTTTCCTCC ACGCCTTCTGTATGAAACAGTTTTTCCTCC ACGCCTTCTGTATGAAACAGTTTTTCCTCC

AdV-NF␬B-Luc

NF␬B

GGGGACTTTCCGCTTGGGGACTTTCCGCT GGGGACTTTCCGCTGGGGACTTTCCGCT GGGGACTTTCCGC

AdEasy-SRE-Luc

Elk1/SRF

CCATATTAGGACATCTAGGATGT CCATATTAGGACATCTAGGATGTCCATATTAGG AGCTAGCCCCATATTAGGACATGCTAGGATGT CCATATTAGGAGCATCTAGGATGTCCCATATTAGG AC

AdEasy-STAT-Luc

STAT

GGTTCCCGTAAATGCATCAGGTTCCCGTAAA TGCATCAGGTTCCCGTAAATGCATCAGG TTCCCGTAAATG

RES-driven luciferase construct

Answer items are indicated by shading or underlining.

3. Results 3.1. Modulated medium from zinc-treated HL-CZ cells induces ICAM-1 protein expression in HUVECs. Our previous study showed that zinc ions alone are sufficient to induce similar levels of ICAM-1 expression as ZnO particles, suggesting that dissolved zinc ions play a major role in the inflammatory effect of ZnO particles on vascular endothelial cells. In this study, we used HL-CZ cells, a human promonocytic leukemia cell line, and the soluble zinc compound Zn(CH3 COO)2 to determine whether zinc ions can induce inflammatory responses in this human promonocytic cell line. HL-CZ cells were treated with different concentrations of Zn(CH3COO)2 (0, 30, 50, 100 and 150 pM) for 6 h, and then conditioned media were collected. HUVEC cells were then treated with conditioned medium for 24 hours. After treatment, HUVEC cell lysates were collected for analysis of ICAM-1 protein expression by immunoblot. The results in Figure 1A showed that ICAM-1 protein expression levels in HUVEC cells were positively correlated with the concentrations of Zn(CH3COO)2 used in HL-CZ treatments. A time-dependent induction of ICAM-1 was also observed in HUVEC cells from conditioned media from HL-CZ cells treated with 150 µM Zn(CH3COO)2 (Figure 1B). Induction of ICAM-1 can be up to 6-7 times. These results suggest that zinc treatments could induce the release of inflammatory cytokines from HL-CZ cells into the culture medium, and the released inflammatory cytokines could activate the expression of ICAM-1 in HUVEC cells. 3.2. Zinc treatments promote the release of chemokines and inflammatory cytokines from HL-CZ cells Since a large number of cytokines have been characterized, how to efficiently identify the expression profiles of multiple cytokines in conditioned media has been complicated. Using the RayBio Human Cytokine Antibody Array 3 to detect secreted/active cytokines, we were able to simultaneously detect 42 cytokine levels in the conditioned medium. The results in fig. 1 shows

that the conditioned medium from HL-CZ cells treated with 150 µM Zn(CH3COO)2 induced maximal level of ICAM-1 expression in HUVEC cells. Media conditioned by such zinc treatments were therefore collected for cytokine analysis with a range of cytokine antibodies. As shown in Table 2, zinc treatment induced a significant release of GRO-␣, IL-6, IL-7, IL-8, and IL-10 with 3.98, 1.92, 1.72, 1.34, and 1.46 times. Although the array detects only 42 cytokines, the results clearly show that zinc treatment induces a significant release of chemokines (eg, GRO-␣ and IL-8), pro-inflammatory cytokines (eg, IL-6 and IL-7), and anti-inflammatory cytokines (eg, IL- 10) from HL-CZ cells. 3.3. Inhibition of NFκB activity by overexpression of IB˛ in HL-CZ cells does not block medium-induced ICAM-1 expression in HUVEC cells, because NF␬B plays a major role in the regulation of zinc-induced ICAM-1 expression in HUVEC cells [14], it was important to additionally ask whether NF␬B also mediates zinc-induced inflammatory cytokine release from HL-CZ cells. By overexpressing I␬B␣ in HL-CZ cells using an adenovirus-mediated expression system, we investigated whether zinc treatment was able to induce I␬B␣ phosphorylation in HL-CZ cells and whether overexpression of I␬B␣ in HL-CZ cells could block media-induced enhanced expression of ICAM-1 in HUVEC cells. As shown in Fig. 2, in uninfected and AdEasyGFP-infected controls, treatment of HL-CZ cells with 150 µM Zn(CH3COO)2 for 6 h caused a 63% knockdown of endogenous I␬B␣. Phosphorylation of I␬B␣ was barely detectable, probably due to rapid polyubiquitination and subsequent degradation of phosphorylated I␬B␣ by the 26S proteasome [29]. In adenovirus-mediated I␬B␣ overexpression experiments, the results showed that inhibition of NF␬B activity by overexpression of I␬B␣ in HL-CZ cells did not prevent conditioned medium-induced ICAM-1 protein expression in HUVEC cells. Meanwhile, zinc treatments enhanced the phosphorylation of exogenous I␬B␣ in HLCZ cells. Due to abundant expression of I␬B␣ by adenovirus

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Table 2. Analysis of the effect of zinc on the release of cytokines from HL-CZ cells by a series of human cytokine antibodies RayBio 3. Cytokines

ENA-78 GCSF GM-CSF GRO GRO-␣ I-309 IL-1␣ IL-1␤ IL-2 IL-3 IL-4 IL-5 IL-6 IL-7 IL-8 IL-10 IL-12 p40p70 IL-13 IL-15 INF-␥ MCP-1 MCP-2 MCP-3 MCSF MDC MIG MIP-1␦ RANTES SCF SDF-1 TARC TGF-␤1 TNF-␣ TNF-␤ EGF IGF-1 MIGF-1 Αγγοειοτγροειοτ τίνη PDGF BB Leptin * **

Inductive bending (processed/unprocessed)

0,540 1,027 0,978 1,854 3,454 1,054 0,799 0,992 1,018 0,550 0,764 0,812 1,930 1,427 1,239 0,790 1. 0 0,998

1,059 1,009 1,484 2,479 5,212 1,186 1,271 1,052 1,447 0,987 2,004 1,735 2,028 1,868 1,348 1,3101 ,223 1,461 1,503 1,168 0 .906 1.099 1.032 0.997 0.864 0.918 0.988 1.017 0.847 1.009 0.920 1.932 0.932 0.920 1.032.

0,914 0,797 0,712 1,634 3,276 0,871 0,836 0,975 1,009 1,046 1,014 0,795 1,787 1,849 1,431 1,409 1,431 1,409 1,009 1,046 0 58 0,872 0,863 0,820 0,862 0,794 0,844 1,011 0,970 0,900 1,028 0,894 0,61074. 5 0,922

Viciously

p-value for one-sample t-test (test value = 1)

0,838 0,944 1,058 1,989 3,981 1,037 0,969 1,006 1,158 0,861 1,261 1,114 1,915 1,715 1,339 0,459 . 0,228 1,196 0,931 1,063 0,985 0,887 0,856 0,963 0,870 0,813 0,908 0,933 0,918 0,939 0,880 0,89470. 7 0,982

0,268 0,128 0,392 0,438 1,070 0,158 0,262 0,040 0,250 0,271 0,656 0,538 0,121 0,249 0,096 0,1940 0,1940 0,1940. 0,103 0,167 0,201 0,161 0,089 0,079 0,123 0,159 0,107 0,082 0,137 0,048 0,1703 4 0,054

0,404 0,530 0,822 0,060 0,040* 0,725 0,855 0,812 0,388 0,468 0,562 0,749 0,006** 0,038* 0,028*0. 4 0,064 0,621

p < 0,05. p < 0,01.

system, we were able to detect the phosphorylation of I␬B␣. These results suggest that NF-B inhibition alone is not sufficient to completely block inflammatory cytokine release from HL-CZ cells. 3.4. Zinc treatment induces the activation of multiple immune response-related transcription factors in HL-CZ cells. Based on our present results, it is suggested that in addition to NF␬B, multiple immune response-related transcription factors may be involved in zinc-induced cytokine release from HL-CZ cells. In order to confirm this hypothesis, seven recombinant adenoviruses carrying a luciferase reporter gene driven by response elements were established (Table 1). These immune response-related transcription factors include AP-1, C/EBP, CREBP, NFAT, NF␬B, SRF, and STAT [30–32]. Activated transcription factors mediate luciferase expression by binding to their respective response elements. HL-CZ cells were infected with one of the recombinant adenoviruses and then treated with 150 pM Zn(CH3COO)2 for 6 h. As shown in Fig. 3, zinc treatment induced the activation of AP1, C/EBP, CREBP, NFAT, NF␬B, SRF, and STAT by 1.18, 2.90, 2.46, 1.64, 4, 27 , 1.42 and 1, respectively, 42-fold in HL-CZ cells. Between them,

C/EBP, CREBP, NFAT, NF␬B and SRF were significantly activated by zinc treatment. 4. Discussion Zinc is an essential trace element for animals and plays an important role in the regulation of immune function in humans [33]. However, excess zinc can also disrupt the homeostasis of the immune system. Epidemiological studies have found an association between increased zinc in the air and increased asthma morbidity in children [10]. Animal studies have shown that ambient PM2.5 samples with higher levels of metals such as zinc cause an increase in allergic airway disease in mice [34]. Our previous in vitro evidence revealed an important role of ZnO particles or its dissolved zinc ions in modulating the inflammatory responses of vascular endothelial cells [14,15]. This study further demonstrates that zinc ions induce the release of chemokines and inflammatory cytokines from vascular promonocytes, possibly through the activation of several immune response-related transcription factors. In a previous study, we assessed vascular endothelial dysfunction using the expression of ICAM-1, a marker of the inflammatory response. ICAM-1, constantly present in low concentrations in

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Fig. 2. Inhibition of NF␬B activity by overexpression of I␬B␣ in HL-CZ cells does not block conditioned medium-induced ICAM-1 protein expression in HUVEC cells. HL-CZ cells were left uninfected (−) or infected with recombinant adenovirus (ADV), AdEasy-I␬B␣ (I␬B␣), or AdEasy-GFP (GFP). The cells were treated with 150 µM Zn(CH3COO)2 for 6 h. After treatment, HL-CZ cell lysates and conditioned media were collected. HUVEC cells were then treated with conditioned medium for 24 h, and HUVEC cell lysates were collected. Using immunoblot analysis, the expression of I␬B␣ (I␬B␣) and phosphorylation (p-I␬B␣) in HL-CZ cells and the protein levels of ICAM-1 and actin in HUVEC cells were determined.

Fig. 1. Conditioned medium (CM) from zinc-treated HL-CZ cells induces ICAM1 protein expression in HUVEC cells. (A) HL-CZ cells were treated with different concentrations of Zn(CH3COO)2 (0, 30, 50, 100 and 150 pM) for 6 h, and then the conditioned medium was collected. HUVEC cells were treated with conditioned medium for 24 hours. (B) HUVEC cells were also treated with conditioned medium for 6, 12 and 24 hours. conditioned medium was collected from HL-CZ cells treated with 150 pM Zn(CH3COO)2 for 6 h. For C2 controls, HUVEC cells were treated with conditioned medium for 24 hours. A 6-hour cultured HL-CZ medium without zinc treatment was used as conditioned medium. For Cl controls, HUVEC cells were cultured in M199 medium for 24 hours. After treatment, HUVEC cell lysates were analyzed for ICAM-1 and actin protein levels by immunoblot analysis. Representative immunoblots are shown (inset). ICAM-1 and actin protein levels were quantified. The experiment was repeated three times. Data are expressed as relative ICAM-1 protein levels compared to untreated control (C1) and are shown as means ± SD. Differences in ICAM-1 expression were examined using Student's t test with 2000 bootstrap samples (for A, 0 vs. C1 and other CM-treated groups vs. 0, for B, other CM-treated groups vs. C2). * p < 0.05.

leukocytes and endothelial cell membranes, is a ligand for LFA-1 (integrin), a receptor found on leukocytes [35]. Here, we showed that conditioned medium from HL-CZ cells treated with 150 µM Zn(CH3COO)2 for 6 hours induced a maximal level of ICAM1 induction (Figure 1). We therefore collected these conditioned media for cytokine secretion analysis using the RayBio Human Cytokine Antibody Array 3. The kit detects secreted cytokines in conditioned media, providing a more accurate reflection of active cytokine levels. Cytokines or other pro-inflammatory mediators induce the expression of adhesion molecules and facilitate leukocyte adhesion to the vascular endothelium via ICAM-1/LFA-1 binding [36,37]. Adhesion of monocytes to the arterial wall and their subsequent infiltration and differentiation into macrophages is a critical step in the development of atherosclerosis. Cytokine analysis in the conditioned medium clearly showed that zinc treatment induced the release of chemokines and inflammatory cytokines from HL-CZ cells. With a range of cytokine antibodies, GRO antibodies detect CXCL1, CXCL2 and CXCL3. That

The GRO-␣ antibody detects only CXCL1. Zinc treatment induced a significant release of GRO (CXCL1, CXCL2, and CXCL3) and GRO-␣ (CXCL1) by 1.99- and 3.98-fold, respectively (Table 2). Since GRO activity involves three CXC chemokines, it is suggested that GRO␣ may be the main CXCL chemokine secreted by HL-CZ cells. GRO-␣, originally isolated and characterized by growth-stimulating activity in malignant melanoma cells, is a chemoattractant for neutrophils [38,39]. Recently, many novel functions of GRO-␣ have been discovered that are associated with atherosclerosis, angiogenesis and many inflammatory conditions [40]. In addition, zinc treatment also induced a significant release of inflammatory cytokines, including IL-6, IL-7, IL-8, and IL-10. In addition to its known role in mediating the acute phase systemic response, IL-6 has multiple roles in the initiation and maintenance of vascular inflammation [41]. IL-7 has multiple roles in T cells, dendritic cells and bone biology in humans and is involved in chronic inflammation that links stroma and adaptive immunity [42]. IL-8, or CXCL8, is also a CXC chemokine and is primarily responsible for the recruitment of monocytes and neutrophils, a signature

Fig. 3. Zinc treatment promotes the activation of several immune response-related transcription factors in HL-CZ cells. HL-CZ cells were infected with a recombinant virus carrying a transcription factor-mediated luciferase reporter gene. Infected cells were treated with 150␮M Zn(CH3COO)2 for 6 h. After treatment, the luciferase activity of each sample was determined. The experiment was repeated three times. Data are expressed as relative reference activity compared to that of the untreated control and are presented as means ± SD. Differences in reference activity (zinc-treated vs. non-zinc-treated) for each transcription factor were tested by a one-sample t-test. *p < 0.05, **p < 0.01 and ***p < 0.001.

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κύτταρα οξείας φλεγμονώδους απόκρισης [43]. Η IL-10, μια σημαντική κυτοκίνη με αντιφλεγμονώδεις ιδιότητες, παράγεται από ενεργοποιημένα ανοσοκύτταρα, ιδιαίτερα μονοκύτταρα/μακροφάγα και υποσύνολα Τ-λεμφοκυττάρων συμπεριλαμβανομένων των κυττάρων Tr1, Treg και Th1 [44]. Με βάση αυτές τις πληροφορίες, τα παρόντα αποτελέσματα καταδεικνύουν τις πιθανές επιπτώσεις της έκθεσης σε ψευδάργυρο στην διαταραγμένη ομοιόσταση της φλεγμονής μέσω αυτών των φλεγμονωδών κυτοκινών. Ο κύριος ρόλος του NF␬B στη ρύθμιση της επαγόμενης από ψευδάργυρο έκφρασης ICAM-1 σε αγγειακά ενδοθηλιακά κύτταρα έχει αποδειχθεί [14]. Ωστόσο, τα αποτελέσματα στο Σχ. 2 υποδηλώνουν ότι, εκτός από το NF␬B, άλλοι μεταγραφικοί παράγοντες που σχετίζονται με την ανοσοαπόκριση μπορεί επίσης να εμπλέκονται στην επαγόμενη από ψευδάργυρο απελευθέρωση χημειοκίνης/φλεγμονώδους κυτοκίνης από κύτταρα HL-CZ. Πράγματι, μεταξύ των επτά μεταγραφικών παραγόντων που σχετίζονται με την ανοσοαπόκριση που δοκιμάστηκαν, οι C/EBP, CREBP, NFAT, NF␬B και SRF ενεργοποιήθηκαν σημαντικά (Εικ. 3). Επομένως, οι επαγόμενες από τον ψευδάργυρο φλεγμονώδεις αποκρίσεις στα προμονοκύτταρα περιλαμβάνουν μια πιο περίπλοκη ρύθμιση σηματοδότησης από αυτή στα αγγειακά ενδοθηλιακά κύτταρα. Επιπλέον, οι συνθήκες θεραπείας με ψευδάργυρο διάρκειας 6 ωρών χρησιμοποιήθηκαν σε αναλύσεις κυτοκίνης και σε δοκιμασία αναφοράς λουσιφεράσης που προκαλείται από μεταγραφικό παράγοντα που σχετίζεται με την ανοσοαπόκριση. Αυτή η βραχυπρόθεσμη θεραπεία σχεδιάστηκε για να αποφευχθούν οι πιθανές δευτερογενείς φλεγμονώδεις αποκρίσεις. Ωστόσο, δεν μπορούσαμε να αποκλείσουμε το ενδεχόμενο στο μεταξύ. Έτσι, απαιτείται λεπτομερής αποκρυπτογράφηση της ενεργοποίησης χρονικής πορείας των μορίων σηματοδότησης που σχετίζονται με τη φλεγμονή και των παραγόντων μεταγραφής στην ακόλουθη μελέτη. Η παρούσα μελέτη ασχολήθηκε κυρίως με τις πιθανές επιπτώσεις στην ομοιόσταση του αγγειακού ανοσοποιητικού συστήματος από αυτές τις υπερβολικές εκθέσεις ψευδάργυρου, ειδικά από σωματιδιακούς ρύπους του περιβάλλοντος. Εκτός από τα αγγειακά ενδοθηλιακά κύτταρα [14], εδώ δείξαμε περαιτέρω ότι ο ψευδάργυρος προκαλεί απελευθέρωση χημειοκίνης και φλεγμονώδους κυτοκίνης από ανθρώπινα προμονοκύτταρα. Με βάση αυτό και τις προηγούμενες μελέτες μας [14,15], περιγράφεται το πιθανό σενάριο φλεγμονωδών αποκρίσεων που προκαλούνται από ψευδάργυρο από σωματιδιακούς ρύπους του περιβάλλοντος. Πρώτον, τα λεπτά σωματίδια τείνουν να παγιδεύονται στις βαθύτερες πνευμονικές κυψελίδες μέσω της εισπνοής. Δεύτερον, η in situ αποσύνθεση των παγιδευμένων σωματιδίων οδηγεί σε τοπική αύξηση των μεταλλικών ιόντων (όπως ο ψευδάργυρος και το νικέλιο) και έτσι μπορεί να ενεργοποιήσει την πνευμονική φλεγμονή. Τρίτον, τα εξαιρετικά λεπτά σωματίδια και τα διαλυμένα μεταλλικά ιόντα μπορούν να διασχίσουν τον φραγμό του πνευμονικού επιθηλίου και στη συνέχεια να εισέλθουν στην κυκλοφορία του αίματος. Τέλος, η άμεση έκθεση των αγγειακών κυττάρων, συμπεριλαμβανομένων των ενδοθηλιακών κυττάρων και των κυκλοφορούντων κυττάρων του αίματος, σε αυτά τα προφλεγμονώδη ιόντα μετάλλων, όπως τα ιόντα ψευδαργύρου σε αυτή τη μελέτη, προκαλεί αγγειακή φλεγμονή. Αυτές οι μελέτες παρέχουν νέα γνώση για την κατανόηση των μηχανισμών αυτών των φλεγμονωδών ασθενειών που προκαλούνται από σωματιδιακούς ρύπους του περιβάλλοντος. 5. Συμπεράσματα Χρησιμοποιώντας ανθρώπινα προμονοκύτταρα HL-CZ ως in vitro σύστημα, αυτή η μελέτη αποκαλύπτει δύο σημαντικά ευρήματα. Η θεραπεία με ψευδάργυρο προκαλεί απελευθέρωση χημειοκίνης και φλεγμονώδους κυτοκίνης από τα κύτταρα HL-CZ. Η διαδικασία περιλαμβάνει την ενεργοποίηση πολλαπλών μεταγραφικών παραγόντων που σχετίζονται με την ανοσοαπόκριση, συμπεριλαμβανομένων των C/EBP, CREBP, NFAT, NF␬B και SRF. Σύγκρουση συμφερόντων Οι συγγραφείς δηλώνουν ότι δεν υπάρχουν σύγκρουση συμφερόντων. Ευχαριστίες Αυτή η εργασία υποστηρίχθηκε από επιχορηγήσεις από τα Εθνικά Ινστιτούτα Ερευνών Υγείας (EO-099-PP-03, EO-100-PP-03) και το Εθνικό Συμβούλιο Επιστημών (NSC97-2314-B-400-003-MY3) στην Ταϊβάν . Είμαστε ευγνώμονες στον Δρ. Shu-Ching Hsu (Έρευνα Εμβολίων και

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Journal of Hazardous Materials 196 (2011) 342-349

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Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Fremstilling af Fe3−x Lax O4 nanokristalni ferit i deres adsorpcijski kapacitet za Congo Red Lixia Wang a,b, Jianchen Li a, Yingqi Wang a, Lijun Zhao a,∗ a b

Key Laboratory of Automotive Materials (Jilin University), Ministry of Education and School of Materials Science and Engineering, Jilin University, Changchun 130022, China School of Mechanical Science and Engineering, Northeast Petroleum University, Daqing 163318, China

article

i n f o

Article history: Received June 2, 2011 Received in revised form September 8, 2011 Accepted September 9, 2011 Available online September 16, 2011 Keywords: La3+-doped magnetite Adsorption Desorption Wastewater treatment

a b s t r a c t This research aimed to increase the adsorption capacity of magnetite for Congo red (CR) by doping a small amount of La3+ ions into it. The adsorption capacity of nanocrystalline Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10) ferrite for the removal of CR from aqueous solution was carefully evaluated. Compared to uncoated magnetite, the adsorption values increased from 37.4 to 79.1 mg g−1. Experimental results show that it is effective to increase the adsorption capacity of magnetite with doped La3+ ions. Among La3+-doped magnetite, Fe2.95 La0.05 O4 nanoparticles show the highest saturation magnetization and maximum adsorption capacity. The desorption capacity of CR-doped magnetite nanoparticles with La3+ can reach 92% after acetone treatment. In addition, Fe3−x Lax O4 nanoparticles showed a distinct ferromagnetic behavior under an applied magnetic field, which enabled their highly efficient magnetic separation from wastewater. High magnetism was found to facilitate the improvement of their adsorption capacity for similar products. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Dyes and pigments are widely used as colorants. Colored organic waste is generated in industries such as textiles, paper, plastics, leather, food and cosmetics, etc. The total consumption of dyes in the textile industry worldwide is over 10,000 tons per year, and about 100 tons of dyes are released into waste streams from the textile industry every year. [1]. It has been recorded that almost 40,000 colors and pigments have been listed, consisting of more than 7,000 different chemical structures [2]. Such colored wastewater can affect the photosynthetic processes of aquatic plants, reduce the level of oxygen in the water and, in severe cases, lead to the suffocation of aquatic flora and fauna [3]. Dyeing wastewater is a pollutant that contains chemicals that have a toxic effect on microbial populations and can be toxic and carcinogenic to organisms and humans. Congo red (CR) (sodium salt of benzidinediazodis-1-naphthylamino-4-sulfonic acid) is metabolized to benzidine, a known human carcinogen, and exposure to this dye can cause some allergic reactions [4]. Treatment of contaminated CR in wastewater is difficult because the dye is generally found in the sodium salt form, which gives it very good solubility in water. Due to their chemical structure, the colors do not fade when exposed to light, water and many chemicals and were therefore difficult

∗ Corresponding author. Tel.: +86 431 85095878; fax: +86 431 85095876. E-mail address:[email protected](L. Zhao). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.032

they change color when the dyes are released into the aquatic environment. Synthetic dyes are difficult to biodegrade due to their complex aromatic structure, which gives them physical-chemical, thermal and optical stability. The high stability of its structure also hinders biodegradation and photodegradation [5]. Various physical, chemical, physicochemical and biological methods (eg nanofiltration [9], enhanced micellar ultrafiltration [10] and electrochemical methods) have been developed to remove dyes and other colored pollutants from wastewater. To remove CR from aqueous solutions But the adsorption capacity of these adsorbents is not high. Activated carbon for adsorbent is too expensive, and both regeneration and disposal of spent carbon are often very difficult [17] Widespread use of some of these adsorbents is limited due to high cost, difficult disposal and regeneration One of the new developments for the removal of dyes from water or wastewater in recent years is the use of ferrite as an adsorbent [18,19]. However, there are still some practical problems to be solved, such as the incompatible relationship between magnetic properties and sizes. As far as we know, decreasing the particle size will increase the surface disorder of nanoparticles. Therefore, the surface energy will increase with decreasing particle size. However, the saturation magnetization of the magnetic powder decreases with decreasing particle size, which is a disadvantage

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High-resolution transmission electron microscopy (HRTEM, JEOL3010, 300 kV) and energy-dispersive X-ray spectroscopy (EDX, Oxford Instruments INCA Energy TEM 200, 300 kV) were used to characterize the microstructure. Hysteresis loops were measured on a VSM-7300 Vibrating Sample Magnetometer (VSM) (Lakeshore, USA) at room temperature. The IR spectra of the samples were characterized using an FTIR spectrophotometer (NEXUS, 670) on KBr pellets. A UV-vis spectrophotometer was used to determine the CR concentration in the solutions.

Figure 1. Structure of Congo red molecule.

2.4. Adsorption tests of magnetic separation after wastewater treatment. In order to solve this problem, this contribution tries to achieve high surface activity by deformation of the crystal structure, which can be shown by the change of the lattice constant after substitution. In addition, we conducted a study of the relationship between adsorption activity and magnetic properties of Fe3−x Lax O4 nanoparticles. Higher magnetic properties were found to facilitate improved adsorption capacity. 2. Materials and methods

The CR mother solution (1 g L-1) was prepared in deionized water, and the desired dye concentrations were obtained by diluting it with water. A calibration curve for CR was prepared by measuring the absorbance of various predetermined sample concentrations at max = 497 nm using a UV-vis spectrophotometer (CR has a maximum absorbance at a wavelength of 497 nm in a UV-vis spectrophotometer). The amount of adsorbed CR (mg g-1) was calculated from the mass balance equation as given below:

2.1. Adsorber

qe =

Congo red [CR, chemical formula = C32H22N6 Na2O6S2, FW = 696.68, max = 497 nm] is an anionic diazo dye based on benzidine, i.e. a dye with two azo groups. The structure is as shown in Figure 1. An accurately weighed amount of dye was dissolved in double-distilled water to prepare a stock solution (1 g L-1).

where qe is the equilibrium adsorption capacity per gram of dry weight of the adsorbent, mg g−1. CO initial concentration of CR in the solution, mg dm−3; Ce is the final or equilibrium concentration of CR in the solution, mg dm−3. V is the volume of the solution, dm3; and W is the dry weight of the hydrogel grains, g. Let's take CR adsorption as an example. A standard solution with initial concentrations of 30 mg L-1 was prepared. Then, 15 mg of Fe3−x Lax O4 nanoparticles were added to 50 mL of the above solution with stirring. After a certain time, the solid and liquid were separated by a magnet, and UV-vis absorption spectra were used to measure the concentration of CR in the remaining solutions. A standard curve, which was used to convert absorbance data to concentrations for kinetic and equilibrium studies, was plotted to calculate the concentration of each experiment.

2.2. Synthesis of Fe3−x Lax O4 Nanocrystalline Ferrite In a typical experiment, FeSO4 ·7H2O and LaCl3 ·7H2O were dissolved in 20 mL of ethylene glycol (EG) with vigorous stirring to obtain a homogeneous solution, followed by 1.5 g of NaOH. It was added to the solution at room temperature with simultaneous vigorous stirring. The mixtures were stirred vigorously for 30 minutes and then sealed in a Teflon-lined stainless steel autoclave and kept at 200°C for 8 hours. After completion of the reaction, the solid product was collected by magnetic filtration and washed several times with deionized water or absolute ethanol. The final product was dried in a vacuum oven at 100 ◦ C for 6 hours. Black powders were obtained and characterized as Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10). Detailed experimental parameters are listed in Table 1 (from S1 to S4). In addition, the experimental work was carried out in winter, so the room temperature was about 13 ◦ C lower.

(C0 − Ce ) × V W

3. Results and discussion 3.1. Characterization of Fe3−x Lax O4 Fig. Figure 1 shows the XRD patterns of S1 and S3. All diffraction peaks in fig. 1a can be attributed to the surface-directed cubic structure of magnetite according to JCPDS card no. 19-0629, and the grating ˚ Diffraction peaks for S3 show that the constant for S1 is 8.40014 Å. it has the same structure as S1 and no impurities can be detected from it

2.3. Characterization Phases were identified by X-ray diffraction (XRD) with a Rigaku D/max 2500 pc X-ray diffractometer with Cu K ␣ () 1.54156 (Å) radiation at a scan rate of 0.02◦ /1(s), morphologies were characterized by JEOL JSM -6700F with a field emission scanning electron microscope (FESEM) operating at an accelerating voltage of 8.0 kV. Transmission electron microscope (TEM, Philips Tecnai 20, 200 kV), Table 1. Summary of experimental parameters. Samples

FeSO4 7H2O (g)

S1 (Fe3 O4) S2 (Fe2.99 La0.01 O4) S3 (Fe2.95 La0.05 O4) S4 (Fe2.90 La0.10 O4)

0,8341 0,8313 0,8202 0,8063

± ± ± ±

0,0002 0,0002 0,0002 0,0002

LaCl3 7H2 O (g) 0,0000 0,0037 0,0185 0,0371

± ± ± ±

0,0002 0,0002 0,0002 0,0002

NaOH (g) 1.5000 1.5000 1.5000 1.5000

± ± ± ±

0,0002 0,0002 0,0002 0,0002

(1)

sl. 1. XRD uzorci: (a) S1 i (b) S3.

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adsorption capacity (mg g-1)

d c

b 40

0 0

100

time (min) Fig. 2. Adsorption capacity: (a) S1; (b) S2; (c) S4 and (d) S3. (Adsorption conditions for CR: 50 mL 100 mg L−1 dye, adsorbent dose 0.015 g, natural pH, temperature: 13 ◦ C.)

˚ Fig. 1b. Additionally, the lattice constant of S3 is 8.39244 Å. The incorporation of La ions can be at the boundaries of magnetite, shortening the length of the Fe-O bond, therefore the lattice constant S3 is smaller than S1. Similar results were published in the work of Zhao and El-Bahy [20,21]. Strong and sharp peaks indicate that S1 and S3 are well crystallized. By comparing the lattice constants between S1 and S3, we can confirm that the crystal structure of magnetite is slightly distorted by the doping of La3+ ions. 3.2. The effect of the amount of La3+ doping on the adsorption capacity of magnetite After we established Fe3−x Lax O4 ferrite in the pure phase, a series of adsorption experiments was carried out. The adsorption capacity from S1 to S4 for CR is shown in the figure. 2. Their adsorption values

for CR are 37.4, 48, 6, 79.1 and 63.1 mg g-1, respectively. An exciting experimental result was obtained that the doping of La3+ ions favors the increase of the adsorption capacity of magnetite for CR. Especially S3 shows the maximum adsorption capacity. Furthermore, at the beginning of the contact time of approx. After 5 minutes, rapid removal of CR was observed. After 90 minutes, adsorption for CR almost reaches saturation. To study the effect of morphology or particle size of Fe3−x Lax O4 on the adsorption capacity of CR in aqueous solution, SEM photographs are shown in the figure. 3a. S1 consists of octahedral nanoparticles with an edge length of about 10-30 nm. Uniform nanoparticles with a size of about 20 nm can be seen from Fig. 3b. However, irregular shapes and broad size distributions appear with increasing content of doped La ions for S3 and S4 (Fig. 3c and d). Their particle sizes are in the range of 60-200 nm and 80-300 nm. As far as we know, the adsorption capacity of nanopowders increased with decreasing particle size (ie increasing surface area). In this experiment, however, it was found that the adsorption capacity can be improved with doped La3+ ions, without a concomitant decrease in particle size. Additional information on the nanostructure of the samples was obtained using TEM and HRTEM. Let's take S3 as an example. Fig. Fig. 4a shows the TEM image of S3, which is consistent with the above SEM results. The HRTEM image (Fig. 4b) and the corresponding fast Fourier transform (FFT) pattern—boxed in Fig. 4b - shown in Fig. 4c represents a face-directed cubic diffraction spot. The clean lattice fringes can show the high crystallinity of the prepared S3. Furthermore, the dominant exposed planes for S3 are {1 1 1}. The lattice spacing between two adjacent edges that we can observe corresponds to a set of (1 1 1) planes with a lattice spacing of 0.5 nm. The lattice edges are parallel along the entire length, which proves the monocrystalline nature of S3. In addition, EDX analysis (Fig. 4d) showed that S3 is mainly composed of Fe, La and O elements, indicating that La ions were introduced into the magnetite.

sl. 3. SEM slike (a) S1, (b) S2, (c) S3 i (d) S4.

L. Wang i sur. / Journal of Hazardous Materials 196 (2011) 342–349

345

Fig. 4. (a) SEM image, (b) HRTEM image, (c) FFT pattern framed in (b), and (d) EDX pattern of S3.

By comparing S1 with other samples, it is certain that properly doped La3+ ions can effectively increase the adsorption capacity. Among S2, S3 and S4, it can be concluded that La3+ ion concentration and particle size show a combined effect on magnetite adsorption capacity. The appropriate amount of doping and particle size lead to the maximum adsorption capacity of magnetite. The ionic radius (r) of La3+ is ˚ rFe 3+ = 0.64 Å). So much larger than Fe3+ (rLa 3+ = 1.06 A, one substitution makes the lattice change constant. Comparing S1 with S3, the lattice constant for S3 is smaller than that for S1. The lattice constant corresponds to the octahedral distortion height ( [MeO6 ]) of magnetite with by a face-centered cubic structure. Decreasing the lattice constant increases the degree of distortion, therefore, magnetite substituted with La3+ ions is more unstable. and they create surface defects in magnetite. Finally, the unstable state can lead to an increase in surface energy. In order to reduce the surface energy of magnetite, it is prone to CR adsorption on its surface. In addition, the structural mismatch caused by doped La3+ ions should change the surface charge of Fe3O4 [23], which can also lead to CR adsorption on the surface based on the principle of electrostatic attachment 3.4 Effect of initial dye concentrations and contact time on adsorption In order to know the effect of initial dye concentration and contact time during CR removal, four different concentrations (0.030 , 0.050 , 0.080 and 0.100 g L−1) were chosen to investigate the adsorption of CR on the Fe2.95 La0.05 O4 surface. By increasing the initial concentrations of CR from 0.030 to 0.100 g L−1, the amount

CR removal increased from 29.2 to 79.11 mg g-1 as shown in the figure. 5. Very fast adsorption is observed in the first 2-5 minutes, then there is a gradual increase with increasing contact time up to 20-30 minutes, depending on the initial dye concentration. Adsorption then maintains a slight increase in the following time. Therefore, adsorption equilibrium occurs after almost 40 min. Similar results were published for CR adsorption on fly ash rich in calcium in the work of Acemio˘glu [24]. 3.5. Adsorption isotherms Analysis of adsorption isotherms is fundamental for describing the interaction of adsorption molecules with the adsorption surface. Two 80 are usually used to simulate isothermal adsorption

adsorption capacity (mg g-1)

3.3. Mechanism of adsorption

100 mg/l 80 mg/l 50 mg/l 30 mg/l

100

time (min) Fig. 5. Effect of initial dye concentration on CR removal from S3. (Conditions: 50 mL CR, adsorbent dose 0.015 g, natural pH, temperature: 13 ◦ C.)

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Table 2 Adsorption parameter obtained from the adsorption isotherm for S3. Freundich

Langmuir

rF 2

qmax (mg g-1)

rL 2

9,8084 ± 0,0002

2,14 ± 0,01

0,9898 ± 0,0002

107,64 ± 0,01

0,0279 ± 0,002

0,9976 ± 0,0001

0,5444 0,4175 0,3094 0,2639

models, Freundlich [25] and Langmuir [26] isotherms, were chosen to explain the dye-ferrite interaction. The Freundlich adsorption isotherm can be expressed as: log qe = log KF +

1 calendar What n

where KF and n are isothermal Freundlich adsorption constants, which are indicative of the extent of adsorption, i.e. the degree of non-linearity between solution concentration and adsorption. The values of KF and 1/n can be calculated from the intercept and slope of the linear plot between log Ce and log qe. The Langmuir isotherm is expressed as:

log(q1 − qt ) = log q1e −

where KL and C0 are the same as previously defined. The RL value is calculated from the above expression. The nature of the adsorption process is either unfavorable (RL > 1), linear (RL = 1), favorable (0 < RL < 1) or irreversible (RL = 0). The Freundlich isotherm was used to describe heterogeneous systems and reversible adsorption, which is not limited to monolayer formations. Unlike the Freundlich isotherm, the Langmuir isotherm is based on the assumption that the structure of the adsorbent is homogeneous, where all adsorption sites are identical and energetically equivalent. Fig. Fig. 6 presents a plot of experimental data based on the Freundlich and Langmuir isotherm models. Table 2 shows the calculated parameter values of the Freundlich and Langmuir models. Comparison of correlation coefficients (r2) linearized

(7)

b y=0,3334x+0,0093

0,025

-1

1/qe (g mg)

logqe(mgg-1)

(6)

h, q2e and K2 can be obtained by a linear plot of t/qt against t. Sl. Figure 7 is a plot of pseudo-first- and second-order kinetics of CR adsorption on S3. The calculated kinetic parameters are given in table 3. The correlation coefficient of the pseudo-first order model is relatively lower (r1 2 = 0.8527), the calculated value of qe (q1e ) was obtained

r = 0,9976

0,020

1.6

0,015

1,5 1,2 1,3 1,4 1,5 1,6 1,7 1,8 1,9 2,0

0,010 0,01

logCe (mgL-1)

(5)

h = K2 q2e 2

0,030

1.7

0,1 0,1 0,1 0,1

where K2 is the pseudo-second order rate constant (g mg−1 min−1). The initial rate of adsorption, h (mg g−1 min−1) at t → 0 is defined as:

1.8

K1 m 2.303

t 1 t = + qt q2e K2 q2e 2

2,0 1,9

± ± ± ±

where qt is the amount of adsorbed dye per unit of adsorbent (mg g−1) in time t, K1 is the pseudo-first order rate constant (min−1). The adsorption rate constant (K1) was calculated from the log(q1e − qt) versus t curve. Ho and McKay [28] presented pseudo-second-order kinetics as:

(4)

y=0,4668x+0,9916 r2=0,9898

30,0 50,0 80,0 100,0

Adsorption kinetic models were used to interpret experimental data to determine the controlling mechanism of dye adsorption from aqueous solution. Here, a pseudo-first-order model, a pseudo-second-order model, and an intraparticle diffusion model were used to test the dynamic experimental data. Lagergren's first-order pseudokinetic model [27] is given as follows:

(3)

1 (1 + KL C0)

0,0002 0,0002 0,0002 0,0002

3.6. Kinetics of adsorption

where qmax is the maximum amount of adsorption with complete monolayer coverage on the surface of the adsorbent (mg g-1), and KL is the Langmuir constant associated with the adsorption energy (L mg-1). The Langmuir constants KL and qmax can be determined from a linear plot of 1/Ce versus 1/qe. The basic characteristics of the Langmuir isotherm can be expressed by a dimensionless constant called the equilibrium parameter RL defined by the following equation: RL =

± ± ± ±

The form of both equations shows that the Langmuir model fits the experimental equilibrium adsorption data better than the Freundlich model. This indicates a monolayer coverage of the S3 surface with CR molecules. The maximum adsorption capacity (qmax) of S3 beads for CR was 107.64 mg g-1 (Table 2). The RL values obtained here are listed in Table 2. All the RL values for CR adsorption on S3 are in the range of 0.5444–0.2639, indicating that the adsorption process is favorable.

(2)

1 1 1 1 = + qe qmax KL qmax Ce

C0 (mg L-1)

0,02

0,03

0,04

0,05

0,06

-1

1/ce (L mg)

Fig. 6. Adsorption isotherms for CR adsorption on S3 (15 mg adsorbent) (a) Freundlich and (b) Langmuir.

L. Wang i sur. / Journal of Hazardous Materials 196 (2011) 342–349

347

Fig. 7. Adsorption kinetics for CR adsorption on S3 (15 mg adsorbent, initial dye concentration 100 mg L−1, natural pH, test temperature: 13 ◦ C) (a) pseudo-first order and (b) pseudo-second order . In (b), the CR solution is re-introduced, which is mixed with magnetic adsorbents and the adsorbent is separated from the solution with a magnet after reaction (2), i.e. 30 minutes.

Table 3 Adsorption parameters obtained from Fig. 5.

79,11 ± 0,01

Pseudo-first class

Pseudo-second order

K1 (min-1)

q1e (mg g-1 )

r1 2

K2 (g mg−1 min−1)

q2e (mg g-1 )

t (mg g−1 min−1 )

r2 2

0,0246 ± 0,0002

12,22 ± 0,01

0,8527 ± 0,0002

0,0085 ± 0,0002

79,43 ± 0,01

53,62 ± 0,01

0,9994 ± 0,0001

from this equation does not give a reasonable value (table 3), which is much lower than the experimental data (qe,exp ). This result suggests that the adsorption process does not follow a pseudo-first-order kinetic model, which is similar to the result reported for CR adsorption on Australian clay materials [29]. In contrast, the results show an ideal fit to second-order kinetics for the adsorbent with an extremely high r22 = 0.9994 (Figure 7b). The good agreement with this adsorption model is confirmed by the corresponding values of the calculated q2e and the experimental values for the adsorbent. The best fit to pseudo-second-order kinetics indicates that the adsorption mechanism depends on the adsorbent and the adsorbate. CR is an acid dye with a negative charge due to the presence of a sulfonated group (-SO3-Na+). Here, the higher adsorption capacity of CR for S3 is probably due to the doping of La ions, which can increase the positive surface charges of magnetite, so we speculate that electrostatic attraction could be the main adsorption mechanism. Insert in fig. Figure 7b presents a photograph of the behavior of adsorption and magnetic separation. After 2 minutes of adsorption, a pink solution was observed. By further extending the adsorption time to 30 minutes, a colorless solution was obtained. More importantly, simple and rapid separation of adsorbed magnetite with CR from purified water can be achieved using an external magnetic field.

at 1050 cm−1, which corresponds to the C-N bond, is shown only in fig. 9b, which reveals that CR was charged on the surface of S3. This also serves as another evidence of physical adsorption due to physical adsorption on the surface of mineral water. oxyanion 100 80

Desorption (%)

qe,exp (mg g-1)

60 40 20 0 0

100

time (min) Fig. 8. Desorption ratio of charged magnetite nanoparticles with time.

3.7. Desorption Desorption is also crucial for the practical application of magnetic powders in water treatment. The simple desorption method and high desorption efficiency can facilitate cost reduction because the used adsorbent and CR can have recycling potential. The desorption process was carried out by mixing 5 mg of CR-modified S3 with 30 ml of acetone solution and shaking for different periods of time. Fig. 8 is the rate of desorption with time. The desorption efficiency was calculated as Eq. (8) was 92%. Therefore, CR can be desorbed from the loaded nanoparticles by acetone solutions. Desorption ratio (%) =

Amount of adsorbed CR × 100 Amount of adsorbed CR

(8)

FT-IR analysis was also performed to reveal the surface nature of S3, as shown in Fig. 9. The spectra show a broad band at 580 cm−1, which is believed to be related to the stretching vibrations of the tetrahedral groups (Fe3+ –O2 − ) for S3. However, the band

Fig. 9. FT-IR spectra of (a) prepared, (b) CR-absorbed and (c) CR-desorbed S3.

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Table 4. Adsorption capacities of CR paint on different adsorbents.

S3 [3] [15] [15] [16] [29] [30] [31] [32] [32] [32] [33] [34] [35] [36] [37] [38] Current examination

it will retain its hydration shell and will not form a direct chemical bond with the oxide surface [22]. Moreover, it is a clear evidence that CR has been sufficiently removed from the surface of S3 by acetone, because Fig. Figures 9a and c show the same FT-IR spectra. 3.8. Performance evaluation The maximum adsorption capacity (qmax) for S3 nanoparticles on CR calculated from the Langmuir isotherm model is given in Table 4 with the literature values of qmax for other adsorbents for CR adsorption [3,15,16,29-38]. All adsorbents used for CR adsorption have significantly lower qmax values than S3 used in this study, except CS/CNT chitosan hydrogel beads impregnated with carbon nanotubes [3], maghemite nanoparticles [16] and modified chitosan beads with CTAB [3738]. However, the simplicity of the preparation method and the magnetic separation of S3 nanoparticles make them better adsorbents than others for CR adsorption. 3.9. Magnetic properties The magnetic properties of magnetic absorbents directly affect memory. Therefore, excellent magnetic performance is also a key role for a magnetic material as a magnetic absorber. Here, the magnetization from S1 to S4 is evaluated. It is exciting to discover that the magnetic properties of magnetic absorbents influence their adsorption capacity. The room temperature hysteresis loops for S1 to S4 are shown in Fig. 10. In addition, the magnetic parameters of the samples obtained from the hysteresis loops are listed in Table 5. The test results show that La3+ doped ions formed Ms values of magnetite somewhat reduced. It should be recognized that magnetite still retains high Ms values after the addition of La3+ ions, although the lowest Ms value is 81.4 emu/g for S2. Surprisingly, we found that the adsorption capacity of La3+-doped magnetite is proportional to their Ms values and independent of their particle size. This is an important evidence that the excellent magnetism facilitates the increase of the adsorption capacity of similar

0 Magnetism (emu/g)

Chitosan hydrogel beads impregnated with CS/CNT carbon nanotubes 450.40 Wheat seeds 22.73 Rice seeds 14.63 208.33 Maghemite cat root nanoparticles 38.79 4.43 Sugar cane bagasse 35.70 7009 pcs. 66.23 Palm seed shell 7.08 Activated red mud 22.62 Aniline propyl silica xerogel 71.46 Seaweed 352.50 CTAB modified chitosan beads 373.29 CS/CTAB beads 107.64 Fe2.95 La0.05 O4

Relationship

Magnetism (emu/g)

qmax (mg g-1)

Type of adsorbent

-40 -80 -10000

-5000

40 20 0 -20 -40 -200

-100 0 100 Field (vi)

5000

200

10.000

Field (Oe) Fig. 10. Magnetization curve measured at room temperature for Fe3−x Lax O4 ferrites.

network products. Both heavy metal ions and organic matter show weak paramagnetism or antiferromagnetism, so magnetic powders with high Ms will be useful for adsorption of heavy metal ions and organic matter. In addition, high Ms magnetic materials aid in final magnetic separation. Therefore, it is very important to keep the magnetic properties almost constant and to increase the adsorption capacity. 4. Conclusions Ferrite nanoparticles Fe3−x Lax O4 (x = 0, 0.01, 0.05, 0.10) were successfully synthesized by simple solvothermal synthesis in one step. Compared to pure magnetite, La3+-doped magnetite shows a better adsorption capacity. Among the La3+-doped products, the sample (Fe2.95 La0.05 O4) with the highest Ms value also has the highest adsorption capacity. The adsorption capacity of magnetite on CR is not improved by increasing the specific area, but by distorting the crystal structure through La3+ doped ions. Compared to many other adsorbents, Fe3−x Lax O4 nanoparticles have a higher adsorption capacity for CR. It would be a good method to increase the adsorption efficiency of magnetite for CR removal in the wastewater treatment process with La3+ ion doping. Analysis of the adsorption isotherm shows that our adsorption experiment agrees with the Langmuir model. Again, the kinetic adsorption model shows that the adsorption mechanism depends on the adsorbent and the adsorbate. In summary, Fe3−x Lax O4 nanoparticles were a kind of excellent adsorbent due to their high adsorption, desorption and recovery efficiency. Acknowledgments We gratefully acknowledge financial support from the Natural Science Foundation of Jilin Province, China (20101542) and the National Doctoral Foundation (grant number 20100061110019). bibliographical references

Table 5. Magnetic parameters obtained from hysteresis loops. Samples

ms (emu/g)

S1 S2 S3 S4

90,5 81,4 86,2 82,2

± ± ± ±

0,1 0,1 0,1 0,1

κ.(emu/g)

hc (you)

± ± ± ±

156,6 116,5 115,9 105,9

15,6 11,2 14,8 16,4

0,1 0,1 0,1 0,1

± ± ± ±

0,1 0,1 0,1 0,1

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Journal of Hazardous Materials 196 (2011) 350-359

Content is available on SciVerse ScienceDirect

Web stranica Journal of Hazardous Materials: www.elsevier.com/locate/jhazmat

Double hydroxide organic/laminar nanohybrids used for the removal of nonionic pesticides D. Chaara a,1, F. Bruna b, M.A. Ulibarri a , K. Draoui c , C. Barriga a, ∗ , I. Pavlovic a Department of Inorganic Chemistry and Chemical Engineering, University Institute of Fine Chemistry and Nanochemistry (IUQFN), University of Córdoba, Rabanales Campus, Campus Agroment Internationalo , Ceia3 , Edificio Marie Curie, 14071 Córdoba, Spain b Institute of Natural Resources and Agrobiology of Seville (IRNAS), CSIC, Avenida Reina Mercedes 10, Apartado 1052, 41080, Sevilla, Spain c Department of Chimie, LCIEs, Factory Science BP 2121, Tetouan, Morocco

article

i n f o

Article history: Received July 15, 2011 Received in revised form September 7, 2011 Accepted September 9, 2011 Available online September 16, 2011 Keywords: organic/layered double hydroxide Nanohybrid pesticide Adsorption Controlled release

ab s t r a c t The preparation and characterization of double hydroxide organo/lamellar nanohybrids with dodecyl sulfate and sebacate as interlayer anions were investigated in detail. The aim of modifying the layered double hydroxides (LDHs) was to change the hydrophilic nature of the interlayer to a hydrophobic one in order to improve the ability of the nanohybrids to adsorb nonionic pesticides such as alachlor and metochlorate from water. Organ/LDH adsorption experiments were conducted using variable pH values, contact times, and initial pesticide concentrations (adsorption isotherms) to determine the optimal conditions for the intended purpose. Adsorbents and adsorption products are characterized by different physicochemical techniques. The adsorption test showed that there was a significant increase in the adsorption of nonionic herbicides. Based on the results, organo/LDH could be good adsorbents for the removal of alachlor and metolachlor from water. Different organo/LDH complexes were prepared by mechanical mixing and adsorption. The results show that the HTSEB-based complex exhibits controlled release properties that reduce metochlor leaching in the soil column compared to the technical product and other formulations. The release depended on the nature of the adsorbent used to prepare the complex. Therefore, it can be concluded that organics/LDHs could act as a suitable support for the design of slow-release pesticide formulations with the aim of reducing the negative effects resulting from rapid transport losses of the chemical after soil application. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Agricultural pesticides are often detected in natural waters, which has raised concerns for health and environmental protection. They are an important group of organic pollutants whose production and use continue to grow, but which need to be controlled in order to minimize pollution problems. production. Layered double hydroxides (LDHs), also known as hydrotalcites, consist of brookite-like layers containing hydroxides of divalent (MII) and trivalent (MIII) metal ions and have an overall positive charge balanced by a hydrated anion.

∗ Corresponding author. Phone: +34 957 218648; fax: +34 957 218621. E-mail address:[email protected](C. Barriga). 1 Permanent address: Department of Chemistry, LPCIE Laboratory, Faculty of Science, BP 2121, Tetouan, Morocco. 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.034

between layers. These compounds show the general formula [MII (1−x) MIII x (OH)2]x+ (An− )x/n ·mH2O, where An− is the interfering anion. Due to the strong hydration of these inorganic ions, the spaces are hydrophilic and resemble clay minerals. As a result, natural clay minerals as well as LDH show rather weak affinity for most nonionic organic compounds and are rarely used as sorbents for organic compounds [3,4] Under appropriate conditions, inorganic ions in clay minerals and layered double hydroxides can be replaced by organic ions that make the interlayer spaces hydrophobic [5-9]. Therefore, the adsorption capacity of clay minerals and LDH for organic pollutants can be significantly improved by modifying the interlayer. These organoclays and organo/LDH are used in a wide range of areas of organic pollution control [10-14]. To date, there are few studies on LDH modified organic anions, and especially their adsorption properties. The number of publications on organoclay (over 2000) is much higher than the number on organo/LDH (about 50) in the last 10 years (source: Citation Report of Web of Science, June 2011). But recently, scientific interest in organic/LDH is increasing. This gives birth

D. Chaara i sur. / Journal of Hazardous Materials 196 (2011) 350–359

in the growing field of research for hydrotalcites, since the results obtained regarding their use as pesticide adsorbents are promising. Organo/LDHs with the appropriate anion can give interlayers between 2 and 4 nm and can be considered as nanohybrids with hydrophobic properties in the interspace and on the outer surface. The aim of this work was to prepare different nanohybrid intercalated hydrotalcites with two organic anions: dodecyl sulfate (DDS) and sebacate (SEB) and different ratios of Mg/Al = 3 and 2, in order to increase the charge density of the layers and consequently increase . intermediate tallow anion content. These nanohybrids were used to evaluate the removal from water of alachlor and metolachlor, two widely used herbicides with hydrophobic properties. In addition, metolachlor was chosen to investigate the release behavior from the formulation with organo/LDH nanohybrids. 2. Materials and methods 2.1. Organic anions, pesticides and soil The organic anions used for the preparation of organo/LDH nanohybrids containing DDS and SEB were available as soluble sodium salt or acid form from Sigma-Aldrich. The values of log Kow are 5.4 and 1.86 respectively for dodecyl sulfate and sebacic acid. (2-chloro-2,6-diethyl-N-(methoxymethyl)) Alachlor and metolachlor (2-chloro-N-(6-ethyl-o-tolyl)-N-[(1RS)-2-methoxy-1- methylethyl]acetamide) are selective preemergence herbicides belonging to the aniline herbicides. Analytical standards alachlor and metolachlor were purchased from Sigma-Aldrich. The solubility values of alachlor and metolachlor in water at 25 ◦ C are 0.110 g/L and 0.120 g/L, and their log Kow values are 2.9 and 3.45, respectively (data obtained from Scifinder Scholar). Molecular structures of used herbicides and organic anions are shown in Fig. 1. The soil used in the leaching experiments was a fluvisol from the terraces of the Guadalquivir River, Córdoba (Spain). Soil samples (0-20 cm) were taken, air-dried and sieved (2 mm) before use. It contained 660 mg/kg sand, 150 mg/kg silt, 190 mg/kg clay and 3.6 mg/kg organic matter. Soil pH was 8.7 in a 1:2 (w:w) soil:deionized water mixture.

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for alachlor or metolachlor. The amount of adsorbed herbicide (Cs) was calculated from the difference between the initial (CO) and the equilibrium concentration of the solution (Ce). Desorption was performed immediately after adsorption from the highest equilibrium point of the adsorption isotherm and was repeated three times. Adsorption-desorption data were adjusted to the Langmuir equation: Ce = Cs

C1e equation

cm L

(1)

and the logarithmic form of the Freundlich equation: logCs = logKf + nf logCe

(2)

where Cm is the maximum adsorption capacity at monolayer coverage (mmol/g), L (L/mmol) is a constant related to the adsorption energy, and Kf (mmol 1−nf Lnf g−1 ) and nf are Freundlich constants. 2.4. Characterization of adsorbents and adsorption products HTSDS1, HTSDS2 and HTSEB adsorbents and adsorption products were characterized by different physicochemical techniques. Powder X-ray diffraction (PXRD) patterns were recorded on the powder samples at room temperature under air conditions using a Siemens D-5000 instrument with Cu K ␣ radiation. FT-IR spectra were recorded using the KBr disk method on a Perkin Elmer Spectrum One spectrophotometer, and the ATR-FT-IR method was used for alachlor. Elemental chemical analyzes for Mg and Al were carried out by atomic absorption spectrometry on a Perkin Elmer AA-3100 instrument. The amounts of DDS and SEB were calculated from the elemental analysis of S and C, respectively, carried out on Elemental Analyze Eurovector EA 3000. The amount of water in the interlayer was obtained from TG curves recorded on a Setaram Setsys Evolution 16/18 instrument, in air at a heating rate of 5 ◦ C/ min. Scanning electron microscope (SEM) micrographs were taken using a JEOL JSM 6300 instrument. The samples were prepared by depositing a drop of the sample suspension on a Cu sample holder and covered with a layer of Au by sputtering on a Baltec SCD005 apparatus.

2.2. Synthesis of organo/LDH nanohybrids

2.5. Preparation of the organo/LDH-methochlorite complex

Organo/LDH containing DDS and SEB anions and Mg/Al = 3 and 2, respectively, were obtained by coprecipitation method [15] using N2 atmosphere and CO2-free water. Samples are shown as HTDDS1 and HTSEB. For comparison, another organ/LDH with DDS was obtained under the same experimental conditions, but without sediment washing. After removing the supernatant, the tube containing the solid was dried. This organ/LDH was named HTDDS2. Carbonate Mg/Al hydrotalcite (HTCO3) was also prepared by the coprecipitation method [16] for the same purpose.

Four complexes with HTDDS1 and HTSEB and metolachlor were prepared. Two of them are based on adsorption isotherms, shown as HTDDS1–MetoAds and HTSEB–MetoAds, and loaded with 3% herbicide. The other two were prepared by mechanically mixing the components by gently grinding the sorbent and herbicide (in the same ratio as the sorbent complexes) dissolved in acetone and then allowing the solvent to evaporate. These are shown as HTDDS1–MetoM and HTSEB–MetoM. 2.6. Experiment with releasing the bath

2.3. Adsorption and desorption experiments The sorption isotherms of alachlor and metolachlor on HTDDS1, HTDDS2 and HTSEB were obtained by the batch equilibration procedure. Triplicate samples of 20 mg of adsorbent were equilibrated by shaking for 24 hours at room temperature with 30 mL of herbicide solution with initial herbicide concentrations (CO) between 0.1 and 0.35 mmol/L. After equilibration, the supernatants were centrifuged and separated to determine the herbicide concentration by UV-visible spectrophotometry at 265 nm and 220 nm.

The release of metolachlor into water from organo/LDH herbicide complexes was compared with the release of the herbicide as a free (technical) product. For this purpose, 0.18 mg of metolachlor was added to 500 ml of distilled water as an organo/LDH herbicide complex or as a technical product. Experiments were performed as described by Bruna et al. [17]. Herbicide concentration was determined by HPLC using a Waters 1525 chromatograph coupled to a Waters 2996 diode detector and UV detection at 220 nm for metolachlor.

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Fig. 1. Molecular structure of organic anions and herbicides.

2.7. Soil Column Leaching Experiments Leaching experiments were performed as described by Bruna et al. [17]. The calculated pore volume of the columns after saturation was 68 ± 2 mL. The washing experiment was performed in triplicate, and metolachlor concentrations were analyzed by HPLC. 3. Results and discussions

The d(0 0 3) spacing of the organo/LDH adsorbents is also included in Table 1. As expected, the positions of the key reflections for all prepared organo/LDHs are shifted toward smaller reflection angles compared to hydrotalcite carbonate, revealing a broadening of the Na interlayer spacings several harmonics are also observed, indicating an ordered structure. The end-to-end van der Waals length of the dodecyl sulfate anion is estimated to be 2.08 nm, taking into account that the thickness of the LDH layer is 0.48 nm, so that the dodecyl sulfate chains can fit perfectly in the vertical direction and

3.1. Characterization of organo/LDH nanohybrid sorbents 3.1.1. Elemental Analysis The results of the elemental analysis for the nanohybrids used as adsorbents are shown in Table 1 along with other characteristics. The organic anion was determined from the S/Al ratio for samples HTDDS1 and HTDDS2 and from C/Al for sample HTSEB. The S content indicates that the layer charge in HTDDS1 is not balanced by the organic anion alone. The exchange rate in the product was 92% based on anion exchange capacity (AEC). However, in HTDDS2 it was 100% and the amount of S was greater than needed to compensate for the layer charge, indicating the precipitation of sodium dodecyl sulfate, as shown in Table 1. Elemental analysis of C for the HTSEB sample showed a small excess sebacata to compensate for the charge of the layer, which could be considered a precipitated salt as shown in Table 1. The proposed formulas are derived from elemental analysis, assuming that all positive charge is compensated by the maximum possible amount of dodecyl sulfate and sebacata cations for HDDDS and HTSEB. The amount of water was obtained from TG data (not included) and metal content analysis. 3.1.2. PXRD X-ray diffraction patterns of the organo/LDH included in Figs. 2 together with HTCO3 show that they are typical compounds of hydrotalcitol. Reflections for HTCO3 are indexed based on a hexagonal unit cell with parameters a and c of 0.304 nm and 2.34 nm, respectively. Corresponding values of

Fig. 2. PXRD patterns of samples HTDDS1, HTDDS2, HTSEB and HTCO3.

D. Chaara i sur. / Journal of Hazardous Materials 196 (2011) 350–359

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Table 1. Chemical composition of adsorbents, structural data and proposed formulas. A sample

HTDDS1 HTDDS2 HTSEB HTCO3

weight%

Atomic ratio

Lille

does

Mg/Al

Sol

C/Al

11,1 9,0 13,5

3,9 3,3 18,1

– 0,35 – –

4,2 5,9 –“ –

27,1 31,3 18,9

3,2 3,1 1,9 2,7

0,9 1,5 – –

15,7 21,7 5,3 –

d0 0 3 (nm)

d1 1 0 (nm)

Recommended species

2,58 3,84 1,48 0,78

0,152 0,152 0,150 0,152

[Mg0,76 Al0,24 (OH)2](C12H25SO4)0,22 (CO3)0,01 ·0,83H2O [Mg0,76 Al0,24 (OH)2](C12H25SO4)0,24 ·0,79H2Oa [Mg 0,65 Al0,35 (OH)2](C10H16O4)0,175 ·1,36H2 Ob [Mg0,73 Al0,27 (OH)2](CO3)0,135 ·0,64H2O

Excess S corresponds to 0.122 NaC12H25SO4 per FW and wt.% N to 0.05 NaNO3. b Same for HTSEB test.

the chains could be in an all-trans configuration [6] in the HTDDS1 sample with a basal distance of 2.58 nm (Table 1). However, for the HTDDS2 sample, a basal spacing of 3.84 nm was observed, similar to those obtained in previous studies [18,19], which could indicate the existence of a bilayer arrangement with a tilt angle of ˛ = 60.4◦ between the DDS chain and layer surfaces [20,21]. A schematic arrangement of the anions in the modified hydrotalcites (nanohybrids) is included in Figure 3. For the HTSEB sample, a basal spacing of 1.48 nm was observed. This results in a gallery height of 1.0 nm, so the anions can be seen as a vertical monolayer and heading towards the brookite-like layers along the tallow length (≈1.3 nm) [17], the required tilt angle is sin − 1 (1, 0 /1.3) = 50.3◦ . The basal distance was shorter than that reported by Bruna et al. obtained for organohydrotalcite with ratio Mg/Al = 3.2, d003 = 1.58 nm. [17]. This may be due to the difference in bed load and/or treatment with the dried layer in each case [18]. The value of the distance corresponding to the (1 1 0) reflection, included in Table 1, is consistent with the metal ratio, which increases slightly as the charge density decreases. 3.1.3. FT-IR spectroscopy Fig. Figure 4 shows the FT-IR spectra of the organo/LDH samples (HTDDS1 is not included because it is almost equal to HTDDS2). The data reveal a LDH-like structure of all adsorbents with their respective intermediate anions (DDS and SEB) as previously reported [22-25]. 3.1.4. Scanning electron microscopy SEM images, included in FIG. 5, showed that the HTCO3 sample consists of a thin plate crystal of irregular shape and size of 1 wt.%. It can be suggested that during N2 flow metal oxide particles need higher temperatures to spread over the carbon surface compared to CO2, which can react at moderate temperatures to remove heteroatoms (O and H) in wood. Finally, it can be seen from Table 8 that the higher the activation or pyrolysis temperature, the more basic the carbon surface. From the pHPZC, it can be concluded that the carbon surface evolves from a soft acidic surface to a basic surface in accordance with the changes in oxygenated functional groups detected by FTIR and XPS discussed above. In other words, the higher the production temperature, the more hydrophobic behavior of carbon is expected. 3.3.2. SEM Figure 5 shows SEM images of selected AC prepared with CO2 and N2 flow. The SEM images show some white dots that have appeared

Table 6 Summary of carbon functional groups detected by XPS analysis in the O1s region for AC prepared under CO2 flow for 1 hour. Temperature (◦ C)

Metal oxides

350 450 600 700 800 900

× × √ √ √ √

Carbonyl √ √ √ √ √ √

Phenol and ether √ √ √ √ √ √

Co-absorbed oxygen or water × × × × √ √

Table 7 Summary of oxide-carbon groups detected by XPS analysis in the O1s region for AC prepared under N2 flow for 1 hour. Temperature (◦ C)

Metal oxides

carbonyl

350 450 600 700 800 900

× × × × √ √

× × × √ √ √

Phenol and ether √ √ √ √ √ √

Co-absorbed oxygen or water × × × × √ √

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Table 8 BET surface area (SBET) and pHPZC of carbon prepared from Algarrob (Alg) under CO2 and N2 flow for 1 hour and comparison with those obtained from Apamata (Apa). T (◦ C)

SBET CO2 (m2/g)

Alg-350 Alg-450 Alg-600 Alg-700 Alg-800 Alg-900 What-450a What-600a What-700a What-800a What-900a

92 350 870 1038 1167 752 352 426 570 770 548

± ± ± ± ± ± ± ± ± ± ± ±

1 7 17 28 31 20 5 13 14 16 21

pHPZC ACCO2

SBET N2 (m2/g)

5,9 6,1 7,0 7,8 8,3 8,9 6,3 7,2 8,0 8,5 9,1

34 220 497 527 549 471 31 360 388 519 590

± ± ± ± ± ± ± ± ± ± ± ±

1 6 1 2 2 12 5 12 13 15 13

Results obtained for Apamata wood under the same experimental conditions [16].

Fig. 5. SEM images of selected activated carbons. (A) CO2, 350 ◦ C, (B) N2, 350 ◦ C, (C) CO2, 600 ◦ C, (D) N2, 600 ◦ C, (E) CO2, 800 ◦ C and (F) N2, 800 ◦ C.

pHPZC ACN2 5,8 5,9 6,9 7,7 7,8 8,3 6,1 7,1 7,9 8,5 8,9

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Fig. 6. Adsorption-desorption isotherm of N2 AC prepared at 800 ◦ C for 1 hour. (A) Under CO2 flow and (B) under N2 flow.

related to the inorganic composition of wood. AC particles are on the micrometer scale according to the sieving performed. Regardless of the gas and activation temperature, AC showed cellular fibrous morphology. SEM images suggest that the present activated carbons consist of an interconnected framework of channels, consistent with the fact that the precursor consists of a fibrous structure. For example, Fig. 5C shows that under CO2 flow, even at moderate temperatures (600 ◦ C), initial activation occurs, although this temperature is lower than what is usually considered to be the critical temperature for spontaneous activation under CO2 flow (approx. ◦ C ) [29]. Similar trends for the interconnected porous system as a function of temperature were reported for the case of activated fibers obtained from rayon fibers [40] and for activated carbon obtained from almond shells [41]. 3.3.3. AC texture For materials prepared at temperatures higher than 600 ◦ C, N2 carbon adsorption-desorption isotherms prepared under CO2 and N2 flow showed very similar trends characteristic of the micropore structure, as shown in the figure. 6 and figures in the Supplementary Material. Fig. Figure 6 shows the analysis for two selected carbon atoms prepared at 800 ◦ C for 1 hour under the flow of CO2 and N2, marked with ACCO2 and ACN2, respectively. Both isotherms correspond to type I, indicating that the framework consists mainly of micropores. An overview of the BET surfaces (SBET) of AC produced from Algarrob (hard wood) is prepared in Table 8, and the results obtained from Apamata (soft wood) under the same experimental conditions [16] are also shown in Table 8 for comparison. In general, the results obtained from Algarrob and Apamata followed very similar trends.

Όπως αναμενόταν, το εμβαδόν επιφανείας BET του AC που παρασκευάστηκε υπό ροή CO2 ήταν υψηλότερο από το υπό N2. Πρέπει να σημειωθεί ότι ένα μέγιστο στο εμβαδόν επιφάνειας BET ελήφθη στους 800 ◦ C, τόσο υπό ροή CO2 όσο και N2. Αυτή η θερμοκρασία είναι η ίδια με αυτήν που βρήκαμε πριν [16] για την παρασκευή AC από το πριονίδι του Apamate [16] και για την ενεργοποίηση αφρού άνθρακα που λαμβάνεται από την ελεγχόμενη πυρόλυση σακχαρόζης υπό ροή CO2 ή N2 [30]. Παράμετροι πορομετρίας όπως εμβαδόν μικροπόρου (πόρος), όγκος μικροπόρου (όγκος πόρου), συνολικός όγκος ή πόρος (Vtot) και διάμετρος πόρων (Wpore) φαίνονται στους Πίνακες 9 και 10 για το AC που λαμβάνεται υπό ροή CO2 και N2, αντίστοιχα. Μπορεί να φανεί ότι όσο υψηλότερη είναι η θερμοκρασία ενεργοποίησης τόσο μεγαλύτερος είναι ο όγκος των μικροπόρων (porevolume ) και τόσο μεγαλύτερος ο συνολικός όγκος των πόρων του AC. Για την περιοχή μικροπόρων (porearea ), επιτυγχάνεται μέγιστο στους 800 ◦ C σε συμφωνία με την επιφάνεια της BET. Στις περισσότερες περιπτώσεις, η μικροπορώδης περιοχή συμβάλλει με περίπου το 90% της συνολικής επιφάνειας. Επιπλέον, μπορεί να φανεί από τους Πίνακες 9 και 10 ότι όσο υψηλότερη είναι η τελική θερμοκρασία ενεργοποίησης τόσο χαμηλότερο είναι το μέσο πλάτος του πόρου (Wpore ). Για θερμοκρασίες μεταξύ 350 και 450 ◦ C ελήφθησαν μακροπορώδεις και μεσοπορώδεις άνθρακες, αντίστοιχα. ενώ ελήφθησαν μεταξύ 600 και 800 ◦ C μικροπορώδη. Μια προσεκτική ανάλυση του όγκου μεσοπόρων που συντάχθηκε στους Πίνακες 9 και 10 έδειξε ότι παρά το πλαίσιο των υλικών άνθρακα είναι κυρίως μικροπόροι, τα δείγματα που παρασκευάστηκαν με αεριοποίηση με CO2 έδειξαν μεγαλύτερη συνεισφορά στην περιοχή μεσοπόρων από τα δείγματα που παρασκευάστηκαν υπό ροή N2. Αυτό ήταν αναμενόμενο γιατί όπως αναφέρθηκε παραπάνω, το CO2 αντιδρά αποτελεσματικά με τον άνθρακα σε θερμοκρασίες υψηλότερες από 600 ◦ C. Γενικά, η μέση διάμετρος πόρων μειώνεται μονοτονικά με την αύξηση της θερμοκρασίας ενεργοποίησης ή πυρόλυσης. Αυτό θα μπορούσε να είναι η συνέπεια μιας ψευδο-γραφιτοποίησης των φύλλων γραφενίου, παρόλο που η παρούσα μέγιστη θερμοκρασία είναι 900 ◦ C, σαφώς χαμηλότερη από αυτή που απαιτείται για αυτό το φαινόμενο. Η μέση διάμετρος πόρων ήταν χαμηλότερη για το AC που ελήφθη υπό ροή N2 από εκείνα που ελήφθησαν υπό ατμόσφαιρα CO2 σε οποιαδήποτε από τις θερμοκρασίες που μελετήθηκαν. Για παράδειγμα, στους 350 ◦ C, η μέση διάμετρος μακροπόρων στη ροή N2 ήταν περίπου το μισό ˚ Αυτή η τάση είναι η ίδια με αυτή στη ροή CO2 (922 έναντι 1815 A). για τη θερμοκρασία όπου λήφθηκαν μεσοπόροι (450 ◦ C), καθώς και στην περιοχή (600–900 ◦ C) όπου ελήφθησαν μικροπόροι AC. Θα πρέπει να σημειωθεί ότι στους 900 ◦ C το χαμηλότερο μέσο πλάτος πόρων περίπου 6,7 A˚ (Πίνακας 9) και 5,3 A˚ (Πίνακας 10) ελήφθησαν υπό ροές CO2 και N2, αντίστοιχα. Αυτά τα AC μπορούν να ταξινομηθούν ως υπερμικροπορώδη AC που έχουν παρουσιάσει πολλές πιθανές εφαρμογές, όπως πυκνωτή διπλής στρώσης και υλικό ηλεκτροδίου [42], ως διαχωριστική μεμβράνη [43], ως καταλυτική υποστήριξη για την παραγωγή υδρογόνου από την αναμόρφωση ξηρού μεθανίου [20 ,21,44,45], και ως αποτελεσματικό προσροφητικό για την πρόσληψη και αποθήκευση υδρογόνου [46]. 3.4. Γενική συζήτηση Παρουσιάζουμε εδώ ένα πρώτο μέρος μιας μεγάλης έρευνας που δείχνει κάποιες γνώσεις σχετικά με το σχεδιασμό τόσο των υφών όσο και των λειτουργικών ομάδων στην επιφάνεια του AC. Μπορεί να συνοψιστεί ότι σε αυτή την εργασία, ένα υδρόφοβο και υπερμικροπορώδες AC μπορεί να παρασκευαστεί ελέγχοντας τη θερμοκρασία και την ατμόσφαιρα της θερμικής αποδόμησης της βιομάζας των αποβλήτων. Το δυναμικό αυτών των AC, κυρίως στη βιομηχανία και σε εφαρμογές περιβαλλοντικής πράσινης χημείας με καταλυτικές και φωτοκαταλυτικές ετερογενείς αντιδράσεις θα παρουσιαστεί στις επόμενες δύο εργασίες. Πρέπει να σημειωθεί ότι τα υλικά που προέρχονται από άνθρακα που λαμβάνονται από πριονίδι ξύλου περιέχουν χαμηλότερη περιεκτικότητα σε τέφρα από εκείνα που παρασκευάζονται από άλλα λιγνοκυτταρινικά υλικά, όπως τα αγροβιομηχανικά βιολογικά απόβλητα και σαφώς χαμηλότερη από εκείνα που λαμβάνονται από πρόδρομες ουσίες πετρελαίου. Αυτός είναι ένας από τους λόγους για τους οποίους τα υλικά ενεργού άνθρακα που παρασκευάζονται από πριονίδι ξύλου έχουν χρησιμοποιηθεί με επιτυχία σε καταλυτικές ετερογενείς αντιδράσεις [1]. Για παράδειγμα, οι Laine et al. [11,12] έχουν δείξει ότι ο όγκος πόρων των υποστηριγμάτων ενεργού άνθρακα παίζει συνεργιστικό ρόλο κατά

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367

Table 9. Micropore surface (pore surface area), micropore volume (pore volume), mesopore volume (mpore volume), total pore volume (Vtot) and average pore width (Wpore) of AC prepared by activation under CO2 flow for 1 hour. T (◦ C) 350 450 600 700 800 900

– 326,7 624,1 892,0 1003 645,7

volumen pora (cm3/g)b

hipervolumen (cm3/g)b

Vtot (cm3/g)c

˚ c Wpora (A)

Hello

0,030 0,107 0,336 0,478 0,538 0,667

1815 54,5 20,6 20,0 18,6 6,72

– 0,100 0,241 0,411 0,462 0,573

– 0,007 0,095 0,067 0,076 0,094

Obtained by the HJ method. Obtained by the HJ method. Obtained according to the HK method. It cannot be calculated.

activity and selectivity of NiMo catalyst in hydrodesulfurization to thiophene. Our group also showed that the pore size distribution clearly affects the catalytic activity of Ni and NiMo catalysts in ethylene hydrogenation [9] and the kinetics of coke deposition [10]. Our group also showed in several works on the synthesis of activated carbon by physical activation or pyrolysis [16] and chemical activation [17] that the pore size distribution and surface area of activated carbon significantly affect the photoactivity of TiO2 in the photocatalytic detoxification of 4-chlorophenol. Furthermore, we have shown that the textural properties of activated carbon clearly influence the selectivity of the main intermediates detected in aromatic molecules such as phenol and 2,4-dichlorophenoxyacetic acid [13,14] and recently for 4-chlorophenol [47, 48] and the photooxidations of 2- of propanol [49]. In this regard, Figure 7 shows the effect of two carbon atoms prepared at 800 ◦ C for 1 hour on phenol adsorption and the photoefficiency of TiO2 in the photodetoxification of phenol under UV radiation. It can be seen from Fig. 7 that one of the two TiO2-AC binary materials adsorbed higher phenol (after 15 minutes of adsorption in the dark). This enhancement of the phenolic molecules around the photoactive TiO2 increases the photoefficiency of the semiconductor, as also shown in Fig. 7. This enhancement is attributed to the presence of a common contact interface between TiO2 and AC that allows continuous species transfer from AC to the TiO2 surface [49]. Our current efforts are focused on the preparation of hierarchical macro-meso-microporous carbon materials to study the effect of pore size distribution on the selectivity of NiMo catalysts in hydrocracking reactions and to verify the presence of limiting effects on the hydrocracking consistency selectivity of the specific pore size of the substrate pore volume as in the case of zeolites [50, 51]. We believe that the present results on the pyrolysis of hardwood sawdust, since the Algarroba essentially consists of 3 different zones, are a very important discovery and deserve careful study. A better explanation for the present results, where the reaction rates decreased considerably with increasing temperature from 350 to 600 ◦ C, could be due to

Waterfalls 6

kapp

cut (min-1)x10-3

c d

Hello

Fadi (mikromolovi)

a b

turbulens (m2 /g)a

TiO2

TiO2 + AC-CO2

TiO2 + AC-N2

Photocatalysts Fig. 7. Summary of the kinetic results of phenol adsorption in the dark (Phads) and the apparent first-order rate constant (kapp) of phenol photodegradation under UV radiation.

the presence of different chemical structures of raw materials at zero retention time after heating to the specified temperatures (ie 350, 400 and 600 ◦ C). Therefore, estimation of kinetic parameters from modulated thermogravimetric analysis data is needed to better understand the relationship between the effect of each crack zone on the pore size distribution and pore volume of carbon materials. In addition, the effect of lignocellulosic carbon precursor ash and the effect of additives (chemical activators) that can play the role of catalysts for improving textural properties are necessary to clarify the significance of the present results. In this way, our groups have already published initial studies on the effect of the pyrolysis atmosphere [52] and the effect

Table 10. Micropore area (pore), micropore volume (pore volume), mesopore volume (mpore volume), total pore volume (Vtot) and average pore width (Wpore) of AC prepared by activation under N2 flow for 1 hour. T (◦ C) 350 450 600 700 800 900 a b c d

pore (m2/g)a d

(Video) UNTITLED Notebook App for Screenwriters - Review (iPad) ($4.99 NOW)

– 201,7 449,1 494,0 515,4 454,8

Obtained by the HJ method. Obtained by the HJ method. Obtained according to the HK method. It cannot be calculated.

volumen pora (cm3/g)b

hipervolumen (cm3/g)b

Vtot (cm3/g)c

˚ c Wpora (A)

Hello

0,016 0,111 0,190 0,201 0,208 0,231

922 43,3 16,3 15,8 15,8 5,33

– 0,101 0,172 0,188 0,195 0,223

– 0,010 0,018 0,013 0,013 0,008

368

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of chemical additives [53] on the topological organization of carbon materials obtained by the controlled pyrolysis of sucrose.

[18]

4. Conclusions [19]

AC is produced from wood sawdust by physical activation and pyrolysis under CO2 or N2 flow. The maximum BET surface area was obtained at 800 ◦ C, in a CO2 and N2 atmosphere, and then decreased at higher activation temperatures. IR and XPS show that the higher the activation temperature, the more basic the functional groups on the carbon surface. Porometry showed that the higher the activation temperature, the greater the volume of micropores and the greater surface area of micropores AC. The higher the activation temperature, the smaller the pore diameter for obtaining ultra-microporous activated carbon at 900 ◦ C. It can be concluded that the average pore width and surface functionality of AC can be easily controlled, and this feature allows thinking about biomass waste as a possible source for the synthesis of carbon materials with various and possible contemporary applications.

[26] [27]

Recognition

[28]

J. Matoš thanks the Ministry of Science and Technology for financial support.

[29]

[20]

[21]

[22] [23] [24]

[25]

[30]

Appendix A. Supplementary data [31]

Supplementary data related to this article can be found in the online version at doi:10.1016/j.jhazmat.2011.09.046.

[32]

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Development of crystallization, microstructure and properties of glass ceramics based on waste sludge prepared by microwave heating Yu Tian a,b,∗ , Wei Zuo a , Dongdong Chen a a b

School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, China State Key Laboratory of Urban Water Resources and Environment, Harbin Institute of Technology (SKLUWRE, HIT), Harbin 150090, China

article

i n f o

Article history: Received April 19, 2011 Received in revised form September 7, 2011 Accepted September 10, 2011 Available online September 16, 2011 Keywords: Wastewater Glass ceramics Dual reactor microwave oven

a b s t r a c t In this study, a microwave melting reactor (MMR) was designed to improve the microwave absorption of waste sludge for the production of glass ceramics. Differential scanning calorimetry (DSC), X-ray diffraction (XRD) and scanning electron microscopy (SEM) were used to study the crystallization behavior and microstructure of the as-grown glass ceramics. DSC and XRD analyzes revealed that crystallization of the nucleated sample in the range of 900–1000 ◦ C resulted in the formation of two crystalline phases: anorthite and wollastonite. When the crystallization temperature was increased from 900 to 1000◦ C, the square grains of wollastonite were subjected to tensile microstrains, causing the crystals to crack. Al ions partially substituted Si ions and occupied tetrahedral sites, causing the formation of anorthite. The relationship between microwave irradiation and crystal growth was investigated, and the result showed that selective microwave heating suppressed crystal growth, giving clear improvements in the properties of glass ceramics. Glass ceramics showed a flexural strength of 86.5–93.4 MPa, a Vickers microhardness of 6.12–6.54 GPa, and a coefficient of thermal expansion of 5.29–5.75 × 10–6 /◦ C. The best chemical resistance to acid and alkali solutions was 1.6.3. ie 0.41-0.58 mg/cm2, showing excellent resistance to alkaline solution. © 2011 Elsevier B.V. All rights reserved.

1. Introduction One of the biggest environmental problems is the safe disposal of the huge amount of waste sludge that is produced every day in wastewater treatment plants [1]. Among the sewage sludge treatment methods, the production of glass ceramics seems promising for the conversion of sewage sludge into new materials with attractive mechanical and chemical properties [2]. Sewage sludge containing large amounts of CaO, SiO2 and Al2O3 can be a good raw material for the production of glass ceramics. By controlling the initial composition and appropriate heat treatment, different crystalline phases are obtained [3]. They exhibit flexural strength, Vickers microhardness, fracture toughness, chemical resistance, and thermal shock resistance that are superior to glass and, in some cases, traditional ceramics [4,5]. It should be noted that the chemical energy of the organic components of sewage sludge can be recovered during the production of glass ceramics as an auxiliary energy source. That

∗ Autor za korespondenciju na: School of Municipal and Environmental Engineering, Harbin Institute of Technology, Harbin 150090, Kina. Tel.: +86 451 8608 3077/13804589869; faks: +86 451 8628 3077. E-mail:[email protected](Y. Tian). 0304-3894/$ – see cover page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.045

Chemical energy in waste sludge can be recovered during the prepared glass-ceramic process as an auxiliary source of energy, thus reducing CO2 emissions, which is suitable for the Kyoto protocol [6]. Other advantages of this technology are the possibility of immobilizing heavy metal ions (retained in glass frames or encapsulated in the crystallization phase) [7], large volume reduction (between 40 and 90%) and flexibility of treatment. process (which can accept different types of sewage sludge, whether municipal or industrial) [8]. The production of glass ceramics by conventional technology is an energy-intensive process, where the process temperature reaches around 1300 ◦ C, and the required process time is several hours [9]. An economic analysis of a glass-ceramic production system capable of processing 0.5-1.0 tons of sewage sludge per hour has shown that the operating costs of this unit are in the range of $100-420 per hour. tons, including labor, fuel and maintenance [10]. Another critical point in the production of glass ceramics is the difficulty in controlling the size and type distribution of crystals due to the thermal inertia of conventional heating [11]. In order to overcome these disadvantages, microwave heating was developed as an alternative technology for the production of dense structural glass-ceramics, which is characterized by a shorter reaction time, reduced energy consumption and reduced crystal size. It was found that the processing temperature decreased

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Fig. 1. Schematic representation of the construction of the reactor for microwave conditioning: (1) microwave cavity; (2) microwave melting reactor (MMR); (3) waveguide; (4) magnetron; (5) Computer with fuzzy logic algorithm. (6) the representative of the government. (7) infrared thermometer.

from 1300 ◦ C to 1000 ◦ C when glass ceramics were melted from barium aluminosilicate glass in a microwave field [12]. It was also reported that a wear-resistant glass-ceramic was grown from the MgO-Al2O3-TiO2 system in 20 minutes by microwave heating [13]. In addition, more uniform and stronger bonding was observed in microwave-prepared glass-ceramics, indicating that microwave energy suppressed grain growth in the crystalline phase due to the high heating rate and obvious low-temperature crystallization [8]. Sewage sludge is a poor recipient of microwave energy to reach the temperature required for the production of glass ceramics. It has been shown that microwave-induced conditioning is possible if an effective receptor is added to the raw sludge. The temperature of sewage sludge can reach 1200 ◦ C in the microwave field when homogeneously mixed with microwave receivers such as graphite and carbon [9]. However, there are fundamental drawbacks to this method when used to produce glass ceramics. The chemical composition of the samples shows uncontrolled changes due to the addition of the microwave receptor, which leads to poor product properties. In addition, the microwave receptor could not be recovered due to the encapsulation of the silicate matrix in the glass ceramic, which increases the operational costs of the process. Attempts called “hybrid microwave sintering” have also been made to solidify around the sample directly to first heat the material to room temperature [14]. However, the temperature of the sewage sludge could not reach high enough due to significant loss of reflection at the interface between the microwave receiver layer and the surrounding air. In order to solve these problems, a new microwave melting reactor (MMR) was designed in this study to produce glass ceramics

from sewage sludge. In MMR, the microwave absorption of sewage sludge can be improved by the double-layer structure, and the required temperature can be reached in a very short time, usually a few minutes. A transparent layer for waves is introduced in the MMR system to reduce the reflection coefficient of the interface between the beam and the MMR. Another important property of the powder was low thermal conductivity, which could give the sample a good quality of thermal insulation. The two-layer structure in MMR ensures a uniform distribution of temperature and electromagnetic field on the samples, which favors the production of glass ceramics with the desired qualities. Additional studies presented in this paper focused on: (1) investigating the effect of the heat treatment program on the crystallization behavior and microstructure of microwave-prepared glass ceramics, (2) defining the evolution of crystallization in the microwave field that was almost never found using conventional procedures, and (3) to gain an insight into the chemical and physical properties of glass ceramics produced by microwave oven in comparison with that obtained by the conventional process. 2. Experimental 2.1. A 2.45 GHz microwave oven designed by MMR, which consisted of a multipurpose rectangular cavity, an infinitely adjustable power supply (0.50–2.7 kW), a temperature control system, and a microwave melting reactor, was used for experimental microwave heating . As shown in Fig. 1, the microwave melting reactor (MMR) consisted of a wave-absorbing layer and a wave-transparent layer. That

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The transparent wave layer is a surface layer that plays an important role in preventing the loss of reflection of the incident wave at the front between the reactor and the air. The wave-absorbing layer below it absorbed the incident wave that was transmitted through the transparent layer and converted the electromagnetic energy into thermal energy. In terms of optimizing MMR performance, the material properties and thickness of each layer are the most important parameters for designing the reactor structure. Activated carbon, a well-known microwave receiver, is loaded in the microwave absorption layer. The material loaded in the microwave transparent layer is selected according to the expression for the reflection coefficient as given in Eq. (1) [15]:

2 - 1 +

(1)

Proximate analysis (wt.%) Aa

Caa, β

Alright

ha, b

i B

I, born

24.50

75,50

39,40

24.43

5,71

4,75

1.18

Content of heavy metals in dry sewage sludge (ppm) Cr

and

143

4,75

138

59,9

700

10.200

38,7

A: ash content. V: content of volatile substances. dry base. b Base without ash. c Calculated by difference.

where R is the reflection coefficient, 1 and 2 are the characteristic impedances for air and material filled with a microwave transparent layer. Obviously, the value of 2 should be close to the value of 1 in order to reduce the reflection coefficient of the incident wave. Based on the results of our preliminary experiments, iron oxides mixed with aluminum oxides (Fe2O3/Al2O3 = 1:1) were used as filling materials for the microwave transparent layer. Before placing in the transparent microwave bed, the Fe2O3 and Al2O3 grains were ground using a mill to obtain a mixed powder with a particle size of ≤75.0 µm. The powder mixture had properties of lower resistance and higher microwave transmission speed, which reduced the reflection coefficient of the air-MMR interface. Another important property of the powder is its low thermal conductivity, which gives the sample a good quality of thermal insulation. The thickness of the microwave transmission layer and the microwave absorption layer is determined by the penetration depth (DE, depth of penetration of microwave energy into the material). DE can be calculated by Fresnel's formula [8]: DE =

Table 1. Chemical properties of sewage sludge.

0 √ εr tgi

(2)

where 0 is the length of the electromagnetic wave in vacuum, εr is the dielectric constant of the material, and tgı is the dielectric loss tangent. According to the calculation of equation (2), the optimal thicknesses of the microwave transmission layer and the microwave absorption layer were determined to be 2 mm. Production of glass from sewage sludge was carried out to investigate the behavior of MMR under microwave heating. It was observed that the sample temperature required for glass production (1300 ◦ C) was reached and the mother glass was successfully produced in this reactor. 2.2. Preparation of mother glass The sewage sludge used in the experiments was collected from a municipal wastewater treatment plant in Harbin, China. Selected chemical properties and heavy metal content of this sludge are listed in Table 1. Dewatered sewage sludge (moisture content was 79.8%) containing small hard particles was crushed in a mortar and then heated to 1000 ◦ C while the sludge samples do not reach a constant weight to remove volatile components. Sewage sludge must be mixed with additives to lower its melting point from a melting point of about 1500 ◦ C. In our experiment, CaO and waste glass were used as effective additives. The chemical compositions of waste sludge, waste glass and raw materials were examined by X-ray fluorescence (XRF) spectroscopy, and the results are shown in Table 2, which indicates that the resulting glass should be in the ternary phase system SiO2-CaO-Al2O3. Fig. Fig. 2 shows the phase diagram of the CaO-Al2O3-SiO2 system. The chemical compositions of raw materials for the production of glass ceramics can be found in

wollastonite-anorthite subsystem (shaded area in Figure 2). The batch composition, prepared by mixing 52.0% raw sludge with 21.0% CaO and 21.0% waste glass, was selected based on the eutectic composition (CaO 38.0, Al2O3 20.0 and SiO2 42.0%) [16 ]. Additionally, 6.0 wt.% TiO2 was added to the base glass composition as a nucleating agent. The mixtures obtained above were melted by microwave and conventional processing. In the microwave process, the glasses were prepared by melting the waste sludge in a corundum crucible at a microwave power of 2000 W for 10 minutes and then naturally cooled to room temperature. In the conventional process, the glass was produced by melting the mixture in an aluminum crucible at 1450 ◦ C for 2 hours, and then the melts were preheated to 600 ◦ C to reduce thermal shock. The results of XRD analysis for waste sludge, glass obtained from these two different heating processes are shown in the figure. 3. 2.3. Glass-ceramic production It is important to accurately determine the nucleation and crystal growth temperatures for the efficient conversion of glass into glass-ceramics. Differential scanning calorimetry (DSC) analysis is performed using a calorimeter (STA449C, NETZSCH) with ␣Al2 O3 as a standard sample. The glass powder is heated from room temperature to 1100 ◦ C at a rate of 10 ◦ C/min to reveal the nucleation and crystallization temperatures. According to the DSC results, the heat treatment program for the microwave-produced mother glass should include a nucleation step at 760 ◦ C for 30 minutes followed by a crystal growth step at different temperatures (900 ◦ C, 950 ◦ C and 1000 ◦ C ) . 60 minutes in microwave radiation. For conventional glass, the sample is held at the nucleation temperature (820 ◦ C) for 90 minutes and then heated to the crystallization temperature (1000 ◦ C) for 120 minutes in an electric furnace. Fig. Fig. 4 shows the production processes of glass ceramics by microwave and conventional methods. The types of crystalline phases were characterized by X-rays with Cu K␣ radiation (XRD: P|max-␥␤, Rigaku, Japan). The step length was 0.02◦ with a scanning speed of 5◦/min in the range 10–90◦ (Cu Ka = 1.5418 A). Schemes for the production of glass ceramics with microwave and conventional heating are shown in Fig. 4. 2.4. Methods for evaluating the properties of glass-ceramics Various techniques were used to evaluate the properties of glass and glass-ceramics. The morphology of the crystalline phases produced on heat-treated glass was examined using a scanning electron microscope (SEM, S-4700, HITACHI). The Archimedean method was used to measure the apparent density of glass ceramics. Hardness and fracture toughness were measured using the indentation method using Vickers indentation. Vickers hardness was measured with loads of 100-1000 g with load

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Fig. 2. Phase diagram of the system CaO–Al2 O3 –SiO2. The chemical composition of the sewage sludge corresponds to the eutectic point indicated by hatching (CS, wollastonite, anorthite CAS2).

times in 10 seconds. The flexural strength was obtained by the four-point method with spans of 20 and 40 mm at a crosshead speed of 100 mm/min, according to the American Society for Testing Materials (ASTM) specification E855-90 [17]. The coefficient of thermal expansion (20–400 ◦ C) was measured by TMA at a heating rate of 10 ◦ C/mm in atmosphere. Chemical resistance was measured according to the American Society for Testing Materials (ASTM) code C27988 [18]. First, powder samples with a particle size of 4.75-6.75 mm were prepared. Then, 20.0 g of the powdered sample was immersed in 100 ml of 1 wt% H2SO4 (about 0.10 mol/l) or 1 wt% NaOH (0.25 mol/l) and boiled on a plate for 48 hours. The samples were dehydrated and the acid/alkali resistance was evaluated by measuring the weight loss of the powders. 2.5. Heavy metal leaching test methods Waste sludge and glass ceramic leaching tests were subjected to the Toxicity Characteristic Leaching Method (TCLP) according to the US Environmental Protection Agency [19]. Sludge and glass-ceramic samples are crushed manually (E > F A > B under the same SIE. 3.1.1. Plasma conversion of HCHO In general, VOCs can be removed by plasma discharge via three routes, i.e. direct electron impacts, gas phase radical attacks and ion collisions Results shown in Figure 3 show that HCHO can be removed not only in the plasma zone (system A) but also in the cylinder after the plasma (system B). For SIE 20 J/L, the conversion of HCHO was 36% and 29% for systems A and B. Considering that unstable plasma species, such as energetic electrons and some gas-phase radicals, cannot reach the post-plasma reactor due to millisecond lifetimes [23, 24], HCHO removal in system B can only be attributed to relatively stable (no O3) and/or metastable plasma species (e.g. N2 metastable states). In fact, it has been reported that for the removal of HCHO from plasma, N2 metastable states may be more important than electrons due to their longer lifetime [25]. These excited N2 states contribute to HCHO removal via two possible pathways: direct attack on HCHO molecules and indirect reactions via the O2 dissociation process, as shown in Eq. (5)-(9) [25,26]. N2 (A3 + u) + HCHO → H + HCO + N2

Fig. 3. Effects of specific input energy on HCHO conversion.

(5)

N2 (a) + HCHO → H + HCO + N2

(6)

3 3 N2 (A3 + u) + O2 → O(P) + O(P) + N2

(7)

N2 (a) + O2 → O(3P) + O(3P) + N2

(8)

O + HCHO → OH + HCO

(9)

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where N2 (a ) denotes the three states a1, a1 and ω1, with an average energy of 8.4 eV. while the energy of N2 (A3 + u) is 6.2 eV for level = 0. In the case of system A in this research, apart from the efficient removal of HCHO in the discharge zone, an additional removal of HCHO can be expected, i.e. which also did not react in to the downstream cylinder due to the remaining long-lived species excited by the plasma.

3.1.2. Καταλυτική μετατροπή του HCHO στο πλάσμα Όπως φαίνεται στο Σχ. 3, το HCHO μπορεί να αφαιρεθεί πιο αποτελεσματικά στα υβριδικά συστήματα πλάσματος-καταλύτη (D-F) σε σύγκριση με τις διεργασίες μόνο του πλάσματος (συστήματα Α και Β). Για SIE 20 J/L, η μετατροπή του HCHO ήταν 87%, 76% και 72% για τα συστήματα D, E και F, αντίστοιχα. Αυτό το αποτέλεσμα υποδηλώνει ότι ο κατάντη καταλύτης MnOx/Al2 O3 μπορεί να ενεργοποιηθεί αποτελεσματικά από είδη πλάσματος μεγάλης διάρκειας ζωής για μετατροπή HCHO σε θερμοκρασία δωματίου και οι ετερογενείς αντιδράσεις οξείδωσης πάνω από την επιφάνεια του καταλύτη είναι πολύ πιο σημαντικές για την απομάκρυνση του HCHO από τις ομοιογενείς αντιδράσεις στην εκκένωση ζώνη. Όπως αναφέρθηκε προηγουμένως, όχι μόνο βραχύβια ασταθή αντιδραστικά είδη παράγονται στις εκκενώσεις πλάσματος, ένα κλάσμα ανασυνδυάζεται για να σχηματίσει πιο σταθερά είδη όπως το O3 [21,27]. Η σύγκριση της μετατροπής HCHO στα συστήματα C και F δείχνει ότι αν και το O3 δεν οξειδώνει το HCHO στην αέρια φάση, ξεκινά την απομάκρυνση του HCHO από τον καταλύτη MnOx/Al2O3. Η έρευνα του μηχανισμού καταλυτικής αποσύνθεσης O3 δείχνει ότι η αποσύνθεση του O3 πάνω από τον καταλύτη οξειδίου του μαγγανίου παράγει ατομικό οξυγόνο και υπεροξείδιο ως ενδιάμεσο είδος [28,29]. Αυτά τα εξαιρετικά ενεργά είδη οξυγόνου θα πρέπει να είναι κυρίως υπεύθυνα για την καταλυτική οξείδωση του HCHO στο σύστημα F. Εκτός από το O3, άλλα μακρόβια διεγερμένα από το πλάσμα είδη, τα οποία ευθύνονται για την απομάκρυνση του HCHO στο σύστημα Β, μπορούν επίσης να υπάρχουν στον δεύτερο κύλινδρο του συστήματος Ε. Αυτά τα είδη μπορεί να προκαλέσουν ομοιογενείς και ετερογενείς αντιδράσεις του HCHO, με αποτέλεσμα υψηλότερη μετατροπή HCHO στο σύστημα Ε από ότι στο σύστημα F υπό τις ίδιες συνθήκες. Τα αποτελέσματα των δοκιμών στο Σχ. 3 δείχνουν επίσης ότι το σύστημα επεξεργασίας D συμπεριφέρεται καλύτερα όσον αφορά τη μετατροπή HCHO σε αυτή τη μελέτη, η οποία μπορεί εύκολα να αποδοθεί στην καλύτερη χρήση των ενεργών ειδών που παράγονται από πλάσμα σε μια διαδικασία καταστροφής HCHO δύο σταδίων : Πρώτον, το HCHO δέχτηκε επίθεση από ενεργητικά ηλεκτρόνια και αντιδρώντα είδη στη ζώνη εκκένωσης. Δεύτερον, το HCHO που δεν αντέδρασε από τη ζώνη εκκένωσης απομακρύνθηκε περαιτέρω στο στάδιο μετά το πλάσμα κυρίως μέσω καταλυτικών διεργασιών που ξεκίνησαν από το O3 και επίσης άλλα μακρόβια ενεργά είδη. Η σύγκριση της μετατροπής HCHO στα συστήματα D, E και F δείχνει ότι οι αντιδράσεις καταλυτικής οξείδωσης που ξεκινούν το O3 παίζουν σημαντικό ρόλο στην καταλυτική απομάκρυνση του HCHO από το πλάσμα. Επιπλέον, πρέπει να σημειωθεί ότι για τα τρία υβριδικά συστήματα, η διαφορά στη μετατροπή HCHO δεν είναι σημαντική για ένα SIE μικρότερο από 3 J/L. Ίσως χρειαστεί να σκεφτούμε ότι αυτό το φαινόμενο συμβαίνει για τους ακόλουθους λόγους. Από τη μία πλευρά, η παραγωγή ειδών υψηλής ενέργειας μακρόβιων, όπως το N2 (A3 + u ) και το N2 (a ), είναι πολύ περιορισμένη στο πλάσμα εκκένωσης χαμηλής ενέργειας. Επομένως, η μετατροπή του HCHO στο σύστημα Ε μπορεί να προκύψει μόνο από τη διαδικασία καταλυτικού οζονισμού όπως ακριβώς και στο σύστημα F. Από την άλλη πλευρά, μπορεί να φανεί από το σχήμα 3 ότι οι ετερογενείς αντιδράσεις είναι πολύ πιο σημαντικές από τις ομοιογενείς αντιδράσεις προς Μετατροπή HCHO, ειδικά στην περίπτωση χαμηλού SIE. Αυτό εξηγεί πιθανώς τη μικρή διαφορά στη μετατροπή HCHO μεταξύ των συστημάτων D και E. Ωστόσο, η διαφορά στη μετατροπή HCHO μεταξύ των τριών υβριδικών συστημάτων γίνεται αξιοσημείωτη σε υψηλότερο SIE, με την αυξανόμενη παραγωγή ειδών υψηλής ενέργειας μακράς ζωής και επίσης υψηλότερη μετατροπή HCHO στη ζώνη εκκένωσης.

Fig. 4. Effects of specific energy intake on energy efficiency regarding HCHO conversion.

In summary, HCHO can be removed not only by short-lived active species in the discharge zone, but also by long-lived species other than O3 downstream of the plasma reactor. Compared to plasma-only processes, sequential plasma-catalyst hybrid systems perform much better in HCHO conversion, mainly due to the heterogeneous destruction of HCHO caused by O3 over the post-plasma MnOx/Al2O3 catalyst. 3.2. Energy efficiency Fig. Figure 4 shows the energy efficiency of HCHO conversion as a function of SIE for treatment systems A, B and D-F. It can be seen that energy efficiency decreased with increasing SIE for all five processing systems. Obviously, the concentration of HCHO in the gas stream decreased with increasing HCHO conversion, resulting in a lower collision probability between HCHO molecules and active species, and thus a lower amount of HCHO removed per kWh of energy consumption at higher SIE. Meanwhile, more energy in the plasma was converted into heat, photons and used to create by-products (as shown in Figure 5) with increasing SIE. A higher SIE favors the complete removal of HCHO (Figure 3), but causes a large energy inefficiency of the process. Therefore, the maximum available value for the input power to the plasma reactor will be determined not only by the conversion of HCHO, but also

Fig. 5. Effects of specific input energy on O3 emission.

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ενεργειακή απόδοση. Ωστόσο, σε σύγκριση με τις διεργασίες μόνο με πλάσμα (συστήματα Α και Β), η ενεργειακή απόδοση βελτιώθηκε σημαντικά με την εισαγωγή καταλύτη MnOx/Al2 O3 μετά το πλάσμα (συστήματα D–F), υποδεικνύοντας ότι οι υβριδικές διεργασίες κατάλυσης πλάσματος έχουν υψηλότερη απομάκρυνση HCHO ικανότητα και είναι πιο ελπιδοφόρα για πρακτικές εφαρμογές. Η μέγιστη ενεργειακή απόδοση ήταν 3,1, 3,1 και 2,5 g/kWh για τα συστήματα D, E και F, αντίστοιχα, σε σύγκριση με 0,9 g/kWh για το σύστημα Α και 0,3 g/kWh για το σύστημα Β. 3.3. Σχηματισμός υποπροϊόντων 3.3.1. Όζον Εφόσον η διεργασία NTP λειτουργεί σε μείγματα που μοιάζουν με αέρα, ο σχηματισμός O3, ενός επικίνδυνου υποπροϊόντος εκκένωσης, είναι αναπόφευκτος. Το Σχ. 5 δείχνει τη συγκέντρωση εξόδου Ο3 ως συνάρτηση του SIE για τα συστήματα επεξεργασίας Α και Δ. Στην πραγματικότητα, οι συγκεντρώσεις εξόδου Ο3 μετρήθηκαν τόσο με την παρουσία όσο και με την απουσία HCHO στον αέρα σε αυτή τη μελέτη για όλα τα συστήματα επεξεργασίας. Τα αποτελέσματα αποδεικνύουν ότι τα χαμηλά επίπεδα HCHO στο ρεύμα αερίου καθώς και η εισαγωγή μιας ρυθμιστικής φιάλης δεν επηρεάζουν σχεδόν καθόλου τη συγκέντρωση εξόδου του O3. Όπως φαίνεται από το Σχ. 5, η συγκέντρωση εξόδου του Ο3 αυξήθηκε με την αύξηση του SIE τόσο για τις υβριδικές διεργασίες του πλάσματος μόνο όσο και για την κατάλυση πλάσματος. Σε σύγκριση με το πλάσμα μόνο, ωστόσο, η παρουσία καταλύτη MnOx/Al2 O3 μετά το πλάσμα μείωσε σημαντικά την εκπομπή O3. Για SIE 20 J/L, η συγκέντρωση εξόδου O3 μειώθηκε από 57,2 ppm για το σύστημα Α σε 13,9 ppm για το σύστημα D. Είναι σαφές ότι κυρίως το O3 που προκλήθηκε από την εκκένωση αερίου αποσυντέθηκε καταλυτικά στην επιφάνεια MnOx/Al2O3, παράγοντας πολύ ενεργών ειδών οξυγόνου που παίζουν βασικό ρόλο στην ενισχυμένη απομάκρυνση του HCHO στις υβριδικές διεργασίες κατάλυσης πλάσματος (Εικ. 3). Ωστόσο, πρέπει να σημειωθεί ότι ακόμη και με την παρουσία καταλύτη MnOx /Al2 O3, η εκπομπή O3 (13,9 ppm για SIE 20 J/L) εξακολουθεί να είναι υψηλή. Σε μελλοντική έρευνα, θα ελεγχθεί εάν η ταυτόχρονη καταλυτική απομάκρυνση του HCHO και του O3 μπορεί να βελτιωθεί περαιτέρω με την εισαγωγή καταλυτών που είναι πιο αντιδραστικοί στην αποσύνθεση του O3. 3.3.2. Μυρμηκικό οξύ Το μυρμηκικό οξύ (HCOOH) είναι ένα κοινό ενδιάμεσο που παράγεται κατά τη διαδικασία οξείδωσης HCHO [5,22]. Οι συγκεντρώσεις εξόδου του HCOOH μετρήθηκαν για τα συστήματα επεξεργασίας Α και D σε αυτή τη μελέτη για τη διερεύνηση της επίδρασης του καταλύτη MnOx/Al2O3 στο σχηματισμό υποπροϊόντων αποσύνθεσης. Προκειμένου να ληφθούν μετρήσιμες συγκεντρώσεις HCOOH, χρησιμοποιήθηκε υψηλότερη αρχική συγκέντρωση HCHO 40,9 ± 0,5 ppm. Τα Σχ. 6α και β δείχνουν τη συγκέντρωση εξόδου και την απόδοση του HCOOH ως συναρτήσεις του SIE, αντίστοιχα. Όπως φαίνεται από το Σχ. 6α, η συγκέντρωση HCOOH στην έξοδο του συστήματος Α αυξήθηκε με την αύξηση του SIE, υποδεικνύοντας ότι το HCOOH πράγματι παρήχθη ως υποπροϊόν στην αποσύνθεση του HCHO στο πλάσμα και η απόλυτη παραγωγή του HCOOH αυξήθηκε με την αύξηση του HCHO αφαιρέθηκε σε υψηλότερο SIE. Αντίθετα, η συγκέντρωση εξόδου HCOOH του συστήματος D μειώθηκε γραμμικά με την αύξηση του SIE και ήταν πολύ χαμηλότερη από αυτή του συστήματος Α στο ίδιο SIE. Για SIE 80 J/L, η συγκέντρωση εξόδου HCOOH ήταν 2,0 και 0,1 ppm για τα συστήματα Α και D, αντίστοιχα. Η διαφορά στην παραγωγή HCOOH μεταξύ των δύο συστημάτων υποδηλώνει ότι το HCOOH που παράγεται στη ζώνη εκκένωσης μπορεί να αφαιρεθεί αποτελεσματικά από τον κατάντη καταλύτη MnOx/Al2O3, ειδικά σε υψηλότερο SIE. Η παρουσία καταλύτη MnOx/Al2 O3 μετά το πλάσμα όχι μόνο ενισχύει σημαντικά τη μετατροπή του HCHO (Εικ. 3), αλλά ευνοεί επίσης την εξάλειψη των οργανικών υποπροϊόντων. Επιπλέον, τα αποτελέσματα που παρουσιάζονται στο Σχ. 6β δείχνουν ότι η απόδοση HCOOH μειώθηκε με την αύξηση του SIE και για τα δύο συστήματα Α και Δ. Δεδομένου ότι η απόλυτη απομάκρυνση του HCHO αυξήθηκε με την αύξηση του SIE, η φθίνουσα απόδοση HCOOH στο σύστημα D μπορεί εύκολα να αποδίδεται στη μειούμενη παραγωγή HCOOH, όπως φαίνεται στο

Fig. 6. Effects of specific energy intake on HCOOH production: (a) HCOOH concentration and (b) HCOOH yield.

Fig. 6a. On the other hand, although the production of HCOOH increased with increasing SIE in system A (Fig. 6a), the yield of HCOOH decreased monotonically, suggesting that more removed HCHO tends to undergo further oxidation reactions in the plasma with higher energy output as end products such as CO2 and CO.

4. Conclusions In this work, the roles of different types of plasma in the plasma and catalytic removal of low concentrations of HCHO in the air were experimentally investigated. The main results can be summarized as follows:

(1) Both short-lived and long-lived plasma species (except O3) contribute to the removal of HCHO in the gas phase. (2) O3 does not initiate the removal of HCHO in the gas phase, but initiates the heterogeneous destruction of HCHO over the MnOx/Al2O3 catalyst, which well explains the significantly improved HCHO conversion by combining the plasma with the MnOx/Al2O3 catalyst in series. (3) The best utilization of plasma-generated active species for HCHO destruction can be achieved in a hybrid plasma-catalyst system where HCHO is introduced through the discharge zone and then the catalytic layer, leading to the highest energy efficiency in terms of HCHO conversion. (4) The introduction of the MnOx/Al2 O3 catalyst after the plasma reactor significantly reduces the emission of discharge by-products (O3) and organic intermediates (HCOOH), showing great potential for indoor VOC purification.

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Acknowledgments The authors thank the National High Technology Research and Development Program (863) of China (No. 2010AA064904) for the financial support for this work. Literature [1] T. Salthammer, S. Mentese, R. Marutzky, Formaldehyde in indoor environment, Chem. Rev. 110 (2010) 2536-2572. [2] D.A. Missia, E. Dimitriou, N. Michael, E.I. Tolis, J.G. Bartzis, Indoor exposure of building materials: A field study, Atmos. Surround. 44 (2010) 4388-4395. [3] X.J. Tang, Y. Bai, A. Duong, M.T. Smith, L.Y. Lee, L.P. Zhang, Formaldehyde in China: production, consumption, exposure levels and health effects, Environ. International 35 (2009) 1210-1224. [4] S. Tanada, N. Kawasaki, T. Nakamura, M. Araki, M. Isomura, Removal of formaldehyde with activated carbon containing amino groups, J. Colloid Interface Sci. 214 (1999) 106-108. [5] F. Shiraishi, D. Ohkubo, K. Toyoda, S. Yamaguchi, Decomposition of gaseous formaldehyde in a photocatalytic reactor with a parallel array of light sources. 1. Basic experiment for reactor design, Chem. Meadow. J. 114 (2005) 153-159. [6] C.B. Zhang, H. He, K.I. Tanaka, Perfect catalytic oxidation of formaldehyde over Pt/TiO2 catalyst at room temperature, Catal. Commune. 6 (2005) 211-214. [7] X.F. Tang, J.L. Chen, Y.G. Li, Y. Li, Y.D. Xu, W.J. Shen, Complete oxidation of formaldehyde over Ag/MnOx-CeO2 catalyst, Chem. Meadow. J. 118 (2006) 119-125. [8] K.Y. Li, Y.N. Shen, M. Jia, S.S. Sheng, M.O. Adebayo, H.Y. Zhu, Catalytic combustion of formaldehyde on gold/iron oxide catalysts, Catal. Commune. 9 (2008) 355-361. [9] R.H. Wang, J.H. Li, OMS-2 catalysts for formaldehyde oxidation: effects of Ce and Pt on catalyst structure and performance, Catal. Easy. 131 (2009) 500-505. [10] J. Van Durme, J. Dewulf, W. Sysmans, C. Leys, H. Van Langenhove, Efficient reduction of toluene in indoor air with a catalytic hybrid plasma system, Appl. A cat. B: Environment. 74 (2007) 161-169. [11] M. Schiorlin, E. Marotta, M. Rea, C. Paradisi, Comparison of atmospheric removal of toluene in air from different corona discharges, Environ. Sci. Technol. 43 (2009) 9386-9392. [12] E. Marotta, A. Callea, M. Rea, C. Paradisi, DC corona electrical discharge for air pollution control. Part 1. Efficiency and products of hydrocarbon processing, Environ. Sci. Technol. 41 (2007) 5862-5868. [13] M.B. Chang, C.C. Lee, Destruction of formaldehyde by dielectric barrier discharge plasma, Environ. Sci. Technol. 29 (1995) 181-186. [14] A. Ogata, H. Einaga, H. Kabashima, S. Futamura, S. Kushiyama, H. Kim, Effective combination of non-thermal plasma and catalyst for benzene degradation in air, Appl. A cat. B: Environment. 46 (2003) 87-95.

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Journal of Hazardous Materials 196 (2011) 386-394

Content is available on SciVerse ScienceDirect

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Biologisk nedbrydning af chlorbenzoesyrer af ligninolytiske svampe ˇ Milan Muzikár a,b , Zdena Kˇresinová a,c , Kateˇrina Svobodová a , Alena Filipová a , Monika Cvanˇcarová a,c , a a,∗ a,∗ Kamila

Department of Microbiology, Academy of Sciences of the Czech Republic, v.v.i., Vídenská 1083, CZ-142 20 Prague 4, Czech Republic ˇ Institute of Chemical Technology Prague, Faculty of Food and Biochemical Technology, Technická 5, CZ-160 28 Prague 6, Czech Institute of Studies of the Environment, Faculty of Science, Charles University, Benátská 2, CZ-128 01 Prague 2, Czech Republic

article

i n f o

Article history: Received May 27, 2011 Received in revised form August 23, 2011 Accepted September 10, 2011 Available online September 16, 2011 Keywords: Chlorobenzoic acid Polychlorinated biphenyls Biodegradation White rot fungi Irpex lacteus

ab s t r a c t We investigated the abilities of several promising strains of ligninolytic fungi to degrade 12 mono-, di- and trichlorobenzoic acids (CBA) under wet model conditions and in contaminated soil. Emphasis was also placed on changes in toxicity during degradation, calculated using two variants of the Vibrio scheri luminescence assay. The results show that almost all fungi were able to efficiently degrade CBA in liquid media, with Irpex lacteus, Pycnoporus cinnabarinus and Dichomitus squalens appearing to be the most efficient in terms of the main factors: degradation and removal of toxicity. The analysis of the decomposition products showed that methoxy and hydroxy derivatives were formed along with the reduced forms of the starting acids. The results suggest that more than one mechanism may be involved in the process. In general, the tested fungal strains were able to degrade CBA in soil in the range of 85-99% within 60 days. Analysis of ergosterol showed that active colonization is an important factor in the degradation of CBA by fungi. The most effective strains in terms of degradation were I. lacteus, Pleurotus ostreatus, Bjerkandera adusta in soil, which were also able to actively colonize the soil. However, unlike P. ostreatus and I. lacteus, B. adusta could not significantly reduce the measured toxicity. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Chlorinated organic pollutants are a class of serious environmental pollutants due to their environmental persistence and ecotoxicity. Chlorinated benzoic acids (CBA) are widespread environmental pollutants that are mainly the result of microbial biodegradation of polychlorinated biphenyls (PCBs), reviewed e.g. in Field and Alvarez [1] and some herbicides [2]. CBAs are significantly more soluble than their parent compounds and therefore can enter the aqueous phase from contaminated soil at contaminated sites. Some mono-, di- and tri-CBAs have been shown to cause genomic damage in tobacco plants [3] and to be toxic to aquatic organisms such as ciliates, daphnia, algae and fish [4-6]. Several mono-, di- and trichlorinated isomers have also been shown to have estrogen interfering activity [7]. CBAs represent critical resistant metabolites in the biphenyl pathway during the transformation of bacterial PCBs. Although CBAs have been shown not to be highly toxic to bacteria, significant adverse effects of their presence on bacterial transformation of PCBs have been reported.

∗ Corresponding author. Phone: +420 241062498; fax: +420 241062384. E-mail address:[email protected](T. Cajthaml). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.041

[8,9]. In addition, soil bacteria that co-metabolize PCBs via the main upstream biphenyl pathway tend to accumulate CBA as dead products because they are generally unable to further transform these substrates [ 10 ]. Another major limitation of organic pollutant biodegradation is the fact that bacterial degradation enzymes are usually intracellular and that transport of the pollutant into the bacterial cell is a major limiting step. On the other hand, ligninolytic fungi, with their extracellular enzymes with low substrate specificity, represent a promising alternative for the biodegradation of various aromatic pollutants [11]. The ligninolytic system consists of three main peroxidases: lignin peroxidase (LiP), manganese peroxidase (MnP), multipurpose peroxidases and laccase, which belongs to phenooxidases. Their degradability has been documented, for example, for chlorophenols, polycyclic aromatic hydrocarbons, PCBs, dioxins, furans, endocrine disruptors and others [12-15]. Furthermore, it has been shown that fungi can cleave the aromatic rings of various persistent pollutants [14,16]. In contrast to the number of articles dealing with the degradation of bacterial CBAs, only a few articles have been published describing the possible degradation of these compounds by fungi. Kamei et al. identified 4-CBA acid after degradation of 4,4-dichlorobiphenyl by Phanerochaete sp. MZ 142 and proposed its further transformation through a reductive pathway [14]. Other authors have shown that ortho and meta mono-CBA and benzoic acid (BA)

Mr. Musician i sur. / Journal of Hazardous Materials 196 (2011) 386–394

προκάλεσε σημαντικά τις δραστηριότητες του κυτοχρώματος P-450 από το Phanerochaete chrysosporium [17]. Στη συνέχεια αποδείχθηκε ότι το ΒΑ μετασχηματίζεται από μικροσώματα που περιέχουν P-450 από τον ίδιο μύκητα. Ο ρόλος των μονοοξυγενασών του P-450 διευκρινίστηκε νωρίτερα από άλλους συγγραφείς, όταν το ένζυμο εκφράστηκε ετερολογικά και ανιχνεύθηκαν βενζοϊκό υδροξυλικό και πρωτοκατεχουϊκό οξύ ως προϊόντα αποικοδόμησης του βενζοϊκού [18]. Γενικά, τα δεδομένα που δημοσιεύθηκαν στο έγγραφο της βιβλιογραφίας ήταν η αποτελεσματικότητα των μυκήτων στην αποικοδόμηση των PCB και υποδηλώνουν πιθανή μετατροπή των PCB σε CBA. Τα CBA είναι κρίσιμοι μεταβολίτες στην οδό αποικοδόμησης των βακτηρίων κυρίως, λόγω της υψηλής εξειδίκευσης των μεμονωμένων βακτηριακών ενζύμων. Ως εκ τούτου, είναι λογικό να διερευνηθούν επίσης οι ικανότητες αποικοδόμησης CBA των λιγνινολυτικών μυκήτων, ειδικά όταν αυτοί οι οργανισμοί αντιπροσωπεύουν μια πολλά υποσχόμενη εναλλακτική λύση σε εφαρμογές αποικοδόμησης βακτηριακών PCB. Ο στόχος αυτής της εργασίας ήταν να διερευνήσει τις ικανότητες αρκετών υποσχόμενων λιγνινολυτικών στελεχών μυκήτων να μετασχηματίσουν 12 εκπροσώπους CBA με διάφορους βαθμούς χλωρίωσης (μονο-, δι-, τρι-CBAs). Η απόδοση αποικοδόμησης δοκιμάστηκε σε μοντέλα υγρών θρεπτικών μέσων, όπου παρακολουθούνταν επίσης η παραγωγή προϊόντων αποδόμησης CBA, οι δραστηριότητες των λιγνινολυτικών ενζύμων και οι αλλαγές στην οξεία τοξικότητα. Επιπλέον, η δυνατότητα εφαρμογής των μυκήτων δοκιμάστηκε επίσης σε ένα τεχνητά μολυσμένο έδαφος, όπου παρακολουθήθηκε επίσης η τοξικότητα. 2. Υλικά και μέθοδοι 2.1. Υλικά Πρότυπα και χημικά. 2-CBA; 2,3-CBA; 3,4-CBA; 3,5 CBA; 2,3,5-CBA; 2,4,6-CBA και HPLC εσωτερικό πρότυπο 2,3διχλωροφαινόλης ελήφθησαν από τη Sigma–Aldrich (Steinheim, Γερμανία). 3-CBA; 4-CBA; 2,4-CBA; Το 2,5-CBA και το 2,6-CBA ήταν από τη Merck (Darmstadt, Γερμανία). Το 2,3,6-CBA αγοράστηκε από τη Supelco (Steinheim, Γερμανία). Όλες οι ενώσεις χρησιμοποιήθηκαν χωρίς περαιτέρω καθαρισμό για την παρασκευή μητρικών διαλυμάτων σε διμεθυλοφορμαμίδιο όπως περιγράφεται παρακάτω. Όλοι οι διαλύτες αγοράστηκαν από τη Merck, Γερμανία ή την Chromservis (Πράγα, Δημοκρατία της Τσεχίας) και ήταν p.a. ποιότητα, ποιότητα ανάλυσης ίχνους ή βαθμός κλίσης. Όλες οι χημικές ουσίες που χρησιμοποιήθηκαν για τις βιοχημικές μελέτες ήταν από τη Sigma–Aldrich (Steinheim, Γερμανία). 2.2. Μικροοργανισμοί, προετοιμασία εμβολίων και μέτρηση ενζυμικών δραστηριοτήτων Καλλιέργειες μυκήτων, πειράματα προετοιμασίας και αποδόμησης εμβολίων. Όλα τα λιγνινολυτικά στελέχη μυκήτων που χρησιμοποιήθηκαν σε αυτή τη μελέτη (Irpex lacteus 617/93, Bjerkandera adusta 606/93, Phanerochaete chrysosporium ME 446, Phanerochaete magnoliae CCBAS 134/I, Traatus87S, Pleurotus 167/93, Pycnoporus cinnabarinus CCBAS 595, Dichomitus squalens CCBAS 750) ελήφθησαν από την Culture Collection of Basidiomycetes της Ακαδημίας Επιστημών της Πράγας. Τα μυκητιακά εμβόλια αναπτύχθηκαν υπό σταθερές συνθήκες για 7 ημέρες στους 28 ◦ C σε φιάλες Erlenmeyer των 250 mL που περιείχαν 20 mL είτε σύνθετου μέσου εκχυλίσματος βύνης-γλυκόζης (MEG) είτε ορυκτού μέσου χαμηλής περιεκτικότητας σε άζωτο (LNMM). Το μέσο MEG (pH 5,5) περιείχε 5 g ζωμό εκχυλίσματος βύνης (Oxoid, UK) και 10 g γλυκόζης ανά λίτρο απεσταγμένου νερού και το LNMM περιείχε 2,4 mM τρυγικό διαμμώνιο [19]. Οι καλλιέργειες στη συνέχεια ομογενοποιήθηκαν με το Ultraturrax-T25 (IKA-Labortechnik, Staufen, Γερμανία) και αυτό το εναιώρημα χρησιμοποιήθηκε για ενοφθαλμισμό στα πειράματα αποικοδόμησης. Προσδιορισμός ενζύμου. Το LiP (E.C. 1.11.1.14) προσδιορίστηκε με βερατρυλική αλκοόλη ως υπόστρωμα [20] και το MnP (E.C. 1.11.1.13) προσδιορίστηκε με 2,6-διμεθοξυφαινόλη [21]. Η λακκάση (Lac, E.C. 1.10.3.2) εκτιμήθηκε με

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2,2-αζινοδις-3-αιθυλβενζο-θειαζολιν-6-σουλφονικό οξύ ως υπόστρωμα [22]. Η ανεξάρτητη από μαγγάνιο υπεροξειδάση (MIP) υπολογίστηκε από τη δραστικότητα υπεροξειδάσης της δοκιμασίας MnP που ανιχνεύθηκε απουσία ιόντων Mn2+. Μία μονάδα ενζύμου παρήγαγε 1 ␮mol του προϊόντος της αντίδρασης ανά λεπτό υπό τις συνθήκες της δοκιμασίας σε θερμοκρασία δωματίου. 2.3. Αποικοδόμηση των CBAs σε υγρά μέσα Τα πειράματα αποικοδόμησης στα υγρά μέσα πραγματοποιήθηκαν ως στατικές καλλιέργειες, επωάστηκαν σε φιάλες Erlenmeyer 250 mL σε πέντε παράλληλα πειράματα στους 28 ◦ C. Είκοσι χιλιοστόλιτρα του αντίστοιχου μέσου (MEG ή LNMM) εμβολιάστηκαν με a5 % εναιώρημα ομογενοποιημένου προ-εμβολιασμού (1 mL) του αντίστοιχου στελέχους μυκήτων. Οι καλλιέργειες εμβολιάστηκαν αμέσως με ένα διάλυμα των CBA σε διμεθυλοφορμαμίδιο (100 ␮L). Η τελική ποσότητα κάθε CBA ήταν 200 ␮g ανά φιάλη. Οι έλεγχοι που σκοτώθηκαν από τη θερμότητα πραγματοποιήθηκαν με ανάπτυξη μιας εβδομάδας μυκητιακών καλλιεργειών, οι οποίες θανατώθηκαν σε αυτόκλειστο πριν από την προσθήκη του διαλύματος CBA. Όλες οι καλλιέργειες επωάστηκαν στο σκοτάδι στους 28 ◦ C και συλλέχθηκαν μετά από 7, 14 και 21 ημέρες. 2.4. Επεξεργασία μολυσμένου εδάφους με μύκητες Για την προετοιμασία του πειράματος αποικοδόμησης του εδάφους, δείγματα 1,0 mL ενός εναιωρήματος μυκηλίων από κάθε στέλεχος μύκητα προστέθηκαν σε δοκιμαστικούς σωλήνες 16 cm × 3,5 cm που περιείχαν 10 g σφαιριδίων αχύρου του εμπορίου (ATEA Praha, Πράγα, Τσεχική Δημοκρατία ), το περιεχόμενο υγρασίας του οποίου είχε προηγουμένως ρυθμιστεί στο 70% (w/w) και στη συνέχεια αποστειρώθηκε σε αυτόκλειστο (121 ◦ C, 45 λεπτά). Μετά τον εμβολιασμό, οι καλλιέργειες κλείστηκαν με πώματα από βαμβάκι και στη συνέχεια αναπτύχθηκαν για 14 ημέρες στους 28 ◦ C [23] Το αποικισμένο υπόστρωμα στη συνέχεια καλύφθηκε με ένα στρώμα εδάφους (20 g), το οποίο είχε προηγουμένως καρφωθεί τεχνητά με ένα μείγμα των CBA σε ακετόνη. Οι σχετικοί μάρτυρες παρασκευάστηκαν με τον ίδιο τρόπο όμως χωρίς μυκητιασικό ενοφθαλμισμό. Οι κύριες ιδιότητες του χρησιμοποιούμενου αμμοαργιλώδους εδάφους ήταν οι εξής: συνολικός οργανικός άνθρακας 0,8%, ολικές οργανικές ουσίες 1,4%, pH 5,3, ικανότητα συγκράτησης νερού 31% και η κοκκομετρική σύνθεση ήταν: άμμος 50,9%, λεπτή άμμος 31,2%, λάσπη 10,8 %, άργιλος 7,1%. Το χώμα ξηράνθηκε στον αέρα και κοσκινίστηκε μέσω πλέγματος 2 mm πριν από τη μόλυνση και η τελική συγκέντρωση κάθε CBA στο έδαφος μετά τη μόλυνση ήταν 10 ␮g/g. Τα δείγματα εδάφους στη συνέχεια υγράνθηκαν σε υγρασία 15%. Οι σωλήνες επωάστηκαν στους 28 ◦ C και τα δείγματα συλλέχθηκαν μετά από 30 και 60 ημέρες. Όλοι οι αντίστοιχοι έλεγχοι και τα δείγματα πραγματοποιήθηκαν σε πέντε επαναλήψεις. 2.5. Εκχύλιση και ποσοτικές αναλύσεις CBAs Ολόκληρο το περιεχόμενο κάθε υγρής καλλιέργειας ομογενοποιήθηκε με Ultraturrax και οξινίστηκε σε περίπου pH 2. Στη συνέχεια εκχυλίστηκε με πέντε δόσεις των 10 mL οξικού αιθυλεστέρα, τα εκχυλίσματα ξηράνθηκαν με θειικό νάτριο και συμπυκνώθηκαν χρησιμοποιώντας περιστροφικό εξατμιστή σε τελικό όγκο 10 mL. Οι ανακτήσεις εξόρυξης όλων των CBA ήταν καλύτερες από 95%. Για να καταστεί δυνατή η ανάλυση HPLC, ένα κλάσμα του εκχυλίσματος οξικού αιθυλεστέρα αναμίχθηκε με ακετονιτρίλιο σε αναλογία 1:10 (v/v) και το μείγμα χρησιμοποιήθηκε για ένεση [16]. Τα δείγματα εδάφους υποβλήθηκαν σε εκχύλιση με χρήση συσκευής εξαγωγής Dionex 200 ASE (Palaiseau, Γαλλία). Τα δείγματα εδάφους (3 g) αναμίχθηκαν με θειικό νάτριο πριν από την εκχύλιση (ο/ο) και οι συνθήκες εκχύλισης ήταν: 3 κύκλοι. 150 ◦ C; 10,34 MPa; σύστημα διαλυτών εξάνιο-ακετόνη, 1% οξικό οξύ [24]. Για να αποφευχθεί η εξάτμιση CBA, 500 ␮L DMSO προστέθηκαν στα εκχυλίσματα ως πώμα διαλύτη και τα εκχυλίσματα συμπυκνώθηκαν χρησιμοποιώντας περιστροφικό εξατμιστήρα κενού (60 kPa, 40 ◦ C) σε περίπου 1,5 mL. 50 ␮L εσωτερικού προτύπου (IS, 2,3-διχλωροφαινόλη 0,9 mg/mL σε ACN) προστέθηκαν σε

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κάθε δείγμα και το IS χρησιμοποιήθηκε για τον υπολογισμό των εκχυλισμάτων όγκου. Το μίγμα στη συνέχεια εγχύθηκε απευθείας στην HPLC. Οι ποσοτικές αναλύσεις πραγματοποιήθηκαν χρησιμοποιώντας το σύστημα Alliance Waters (Πράγα, Δημοκρατία της Τσεχίας) εξοπλισμένο με ανιχνευτή PDA και το λογισμικό Empower χρησιμοποιήθηκε για την επεξεργασία δεδομένων. Ο διαχωρισμός του μίγματος CBA πραγματοποιήθηκε σε στήλη XBridge C18 (250 mm x 4,6 mm I.D., μέγεθος σωματιδίων 3,5 ␮m) από την Waters (Πράγα, Δημοκρατία της Τσεχίας). Ο διαχωρισμός διεξήχθη με βαθμίδωση (ο/ο) ακετονιτριλίου (Β) και υδατικού διαλύματος (Α) 0,1% TFA. Το πρόγραμμα κλίσης ήταν ως εξής (min/%B): 0/17; 30/17; 60/34; 70/50. Ο εφαρμοζόμενος ρυθμός ροής ήταν 0,8 mL min−1 και η θερμοκρασία ήταν 35 ◦ C [24]. 2.6. Ποιοτικές αναλύσεις προϊόντων αποδόμησης CBA Πραγματοποιήθηκε ποιοτική ανάλυση των προϊόντων αποδόμησης CBA με το ότι τα ενδιάμεσα προϊόντα αποδόμησης διαχωρίστηκαν και χαρακτηρίστηκαν ή ταυτοποιήθηκαν με αέρια χρωματογραφία-φασματομετρία μάζας (GC–MS; 450-GC, 240-MS ανιχνευτής παγίδας ιόντων, Varian, Walnut Creek, CA). Τα εκχυλίσματα οξικού αιθυλεστέρα εγχύθηκαν τόσο απευθείας χωρίς καμία παραγωγοποίηση όσο και μετά από τριμεθυλσιλυλίωση με Ν,Ο-δις(τριμεθυλσιλυλ)τριφθοροακεταμίδιο (BSTFA, Merck, Γερμανία) και μεθυλίωση με διαζωμεθάνιο [25]. Το όργανο GC ήταν εξοπλισμένο με έναν εγχυτήρα διαχωρισμού/χωρίς διαχωρισμό που διατηρείται στους 240 ◦ C. Η στήλη DB-5MS (Agilent, Πράγα, Τσεχική Δημοκρατία) χρησιμοποιήθηκε για τους διαχωρισμούς (30 m, 0,25 mm I.D., 0,25 mm πάχος φιλμ). Το πρόγραμμα θερμοκρασίας ξεκίνησε στους 60 ◦ C και διατηρήθηκε για 1 λεπτό στη λειτουργία splitless. Στη συνέχεια ο διαχωριστής άνοιξε με αναλογία 1:50. Ο φούρνος θερμάνθηκε στους 120 ◦ C με ρυθμό 25 ◦ C/min με μια επακόλουθη ράμπα θερμοκρασίας στους 240 ◦ C με ρυθμό 2,5 ◦ C/min, όπου αυτή η θερμοκρασία διατηρήθηκε για 20 λεπτά. Ο χρόνος καθυστέρησης του διαλύτη ορίστηκε στα 5 λεπτά και η θερμοκρασία της γραμμής μεταφοράς ορίστηκε στους 240 ◦ C. Τα φάσματα μάζας καταγράφηκαν σε 3 σαρώσεις s−1 υπό κρούση ηλεκτρονίων στα 70 eV και εύρος μάζας 50–450 amu. Το δυναμικό διέγερσης για τον τρόπο παραγωγής ιόντων MS/MS που χρησιμοποιήθηκε ήταν 0,2 V και αυξήθηκε σε 0,8 V για πιο σταθερά ιόντα. Το ακετονιτρίλιο χρησιμοποιήθηκε ως μέσο για χημικό ιονισμό (CI), όπου ο μέγιστος χρόνος ιοντισμού ήταν 2000 και 40 ␮s για την αντίδραση. 2.7. Αναλύσεις εργοστερόλης Η ολική εργοστερόλη εκχυλίστηκε και αναλύθηκε όπως περιγράφηκε προηγουμένως [23]. Εν συντομία, τα δείγματα (0,5 g) υποβλήθηκαν σε υπερήχους με 3 mL 10% ΚΟΗ σε μεθανόλη στους 70 ◦ C για 90 λεπτά. Προστέθηκε απεσταγμένο νερό (1 mL) και τα δείγματα εκχυλίστηκαν τρεις φορές με 2 mL κυκλοεξανίου, εξατμίστηκαν υπό άζωτο, επαναδιαλύθηκαν σε μεθανόλη και αναλύθηκαν ισοκρατικά χρησιμοποιώντας ένα σύστημα HPLC Waters Alliance (Waters Milford, MA) εξοπλισμένο με στήλη LiChroCart γεμάτη με LiChrosphere® 100 RP-18e (250 × 4,0 mm, μέγεθος σωματιδίων 5 ␮m, ˚ εξισορροπημένο με 100% μεθανόλη σε ρυθμό ροής μέγεθος πόρων 100 A) 1 mL min−1. Η εργοστερόλη ανιχνεύθηκε στα 282 nm και προσδιορίστηκε ποσοτικά με καμπύλη βαθμονόμησης 5 σημείων σε γραμμικό εύρος από 0,5 έως 50,0 ␮g/mL. 2.8. Δοκιμασία τοξικότητας Τα φωτοβακτήρια Vibrio fischeri (στέλεχος NRRL-B-11177), που χρησιμοποιήθηκαν για όλες τις δοκιμές τοξικότητας, αγοράστηκαν λυοφιλοποιημένα από τον προμηθευτή Ing. Musial (Τσεχία). Τα λυοφιλοποιημένα βακτήρια επανυδατώθηκαν και σταθεροποιήθηκαν σε διάλυμα NaCl 2% (w/v) στους 15 ◦ C για 1 ώρα σύμφωνα με την τυπική διαδικασία ISO 2007 [26]. Πραγματοποιήθηκε δοκιμή οξείας τοξικότητας δειγμάτων μετά από αποικοδόμηση σε υγρά μέσα χρησιμοποιώντας τα αντίστοιχα εκχυλίσματα οξικού αιθυλεστέρα. Δείγματα των εκχυλισμάτων (0,5 mL) εξατμίστηκαν μέχρι ξηρού και διαλύθηκαν ξανά σε διμεθυλοσουλφοξείδιο, το οποίο εφαρμόστηκε απευθείας στη δοκιμή (2% DMSO στο μίγμα της αντίδρασης).

The amount of dimethyl sulfoxide varied between media due to different sensitivities of the assay to the media matrix (see below). Three replicates were used for each sample to perform the ecotoxicity test. Luminescence measurements were performed with a Lumino M90a photometer (ZD Dolní Újezd, Czech Republic) at a temperature of 15 ± 0.2 ◦ C. Bioluminescence inhibition was recorded after a 15-minute exposure. The toxicity of soil samples was measured using the Flash kinetic test using luminescent bacteria [27,28]. The samples were prepared by weighing 1.5 g of dry soil and 6 ml of 2% (w/v) NaCl solution. The sample suspension was continuously stirred and 0.5 ml was placed in the measuring cell. The contents of the measuring cell were continuously mixed with a customized LUMINO M90a photometer and 0.5 ml of the bacterial solution was dosed into the sample. The signal was permanently recorded for 60 seconds. Light inhibition was calculated as the difference between the height of the peak observed immediately after the addition of bacteria to the sample and the luminescence intensity after a contact time of 60 s.

3. Results and discussion 3.1. Degradation of CBA in liquid cultures The representatives of mono-, di- and tri-CBA tested in this study were used at a relatively high concentration of 10 ␮g/ml. The fact that the compounds are partially soluble in water and their acute toxic properties were confirmed by observing the growth of mushroom biomass. In general, the fungal strains were partially affected by CBA and their biomass reached about 50–70% compared to non-toxic controls (data not shown). This finding is consistent with the observation of Dittmann et al. who tested mycelial growth of fungal strains in two liquid media after addition of different concentrations of 3-CBA [29]. Contrary to this observation, the fungal strains in our study were very efficient in degrading CBA in liquid media. The time course of individual CBA degradation in both media is shown in Tables 1 and 2. The results clearly show that all tested strains were at least partially capable of transforming CBA. I. lacteus, P. cinnabarinus and D. squalens were found to be the most efficient decomposers in complex MEG media. P. cinnabarinus and D. squalens were able to decompose approx. 78% and 73% of total CBA, respectively, while I. lacteus degraded 92% of total CBA in complex medium compared to heat-killed controls. In particular, I. lacteus removed all CBA from the medium except 2,6-CBA and 2,3,6-CBA, which were degraded to about 50% of the original amount, while P. cinnabarinus did not significantly transform 2,6 . - CBA, 2,3,6-CBA, 2,4,6-CBA and D. squalens did not significantly degrade 2,3,6-CBA (ANOVA, P = 0.05). B. adusta showed the worst degradation ability in both liquid media. The most effective strains on the LNNM medium were again I. lacteus, P. cinnabarinus and D. squalens. All three of the most effective strains were able to transform CBA with percentage removal ranging from 76% to 77%. Again, 2,6-CBA, 2,3,6-CBA and 2,4,6-CBA appear to be the most persistent compounds, with D. squalens not significantly degrading 2,6-CBA and 2,4,6-CBA ; P. cinnabarinus – 2,6-CBA, 2,3,6-CBA and 2,4,6-CBA; I. lacteus – 2,6-CBA and 2,3,6-CBA. These results suggest a possible link between substituted ortho- and parasites and persistence according to the fungal degradation mechanism. Several journal publications have addressed the fungal transformation of CBA. The aforementioned work by Dittmann et al. also involves the degradation of 3-CBA by P. chrysosporium, P. ostreatus, Heterobasidion annosum and two other ectomycorrhizal fungi [29]. However, in contrast to our results, the authors observed only limited degradation of the link in the region of a few

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Table 1 Restmøynger af CBA i varmedræbte kontroller og nakon incubacije nakon testiranih svampestammer u LNNM-mediju (ND: ikke påvist). Mængde CBA (␮g pr. flaske) LNNM – 7 dana Kontrola B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor LNNM – 14 dana Kontrola B. adusta Iqualens D. lacteus P chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor LNNM – 21 dan Kontrol B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T.

2-CBA

3-CBA

4-CBA

2,3-CBA

2,4-CBA

2,5-CBA

2,6-CBA

3,4-CBA

3,5-CBA

2,3,5-CBA

2,3,6-CBA

2.4.6-CBA

196 ± 4 156 ± 5 ND ND ND 27 ± 2 156 ± 9 ND 73 ± 39

207 ± 4 53 ± 12 33 ± 4 19 ± 6 ND ND 28 ± 3 ND 25 ± 1

207 ± 2 41 ± 34 ND ND ND ND ND ND ND

195 ± 3 174 ± 5 ND 45 ± 7 94 ± 14 133 ± 2 162 ± 8 ND 25 ± 3

207 ± 2 156 ± 31 ND 66 ± 2 ND 74 ± 4 105 ± 17 ND 104 ± 8

193 ± 3 141 ± 11 ND 85 ± 5 92 ± 15 126 ± 6 167 ± 8 ND 46 ± 3

205 ± 2 197 ± 8 211 ± 15 179 ± 18 196 ± 5 189 ± 3 195 ± 5 190 ± 8 242 ± 11

182 ± 8 64 ± 19 ND ND ND ND 87 ± 60 ND ND

196 ± 1 77 ± 28 ND ND ND 165 ± 2 55 ± 1 ND ND

165 ± 2 161 ± 2 ND 153 ± 2 136 ± 10 ND 124 ± 46 ND ND

199 ± 1 185 ± 7 ND 176 ± 21 177 ± 7 184 ± 8 172 ± 14 176 ± 12 206 ± 14

185 ± 0 176 ± 8 207 ± 6 128 ± 11 171 ± 6 180 ± 4 163 ± 13 173 ± 7 140 ± 4

170 ± 15 124 ± 5 ND ND ND 105 ± 1 ND 66 ± 13

173 ± 14 45 ± 2 56 ± 4 ND ND ND 47 ± 4

184 ± 21 ND ND ND ND ND ND ND ND

168 ± 19 157 ± 22 126 ± 8 13 ± 2 25 ± 1 100 ± 11 132 ± 1 ND ND

208 ± 17 85 ± 5 ND ND ND 60 ± 19 15 ± 2 ND ND

167 ± 18 114 ± 7 ND ND 40 ± 1 77 ± 13 159 ± 3 ND ND

181 ± 16 198 ± 14 189 ± 11 173 ± 7 188 ± 12 182 ± 15 193 ± 11 184 ± 6 186 ± 34

156 ± 17 46 ± 0 ND ND ND ND ND ND ND

165 ± 19 ND ND ND ND 136 ± 23 ND ND ND

138 ± 19 145 ± 5 ND 103 ± 3 92 ± 0 ND 114 ± 6 ND ND

169 ± 17 187 ± 7 ND 171 ± 9 175 ± 12 153 ± 32 188 ± 5 163 ± 8 183 ± 3

159 ± 21 179 ± 10 175 ± 11 95 ± 4 165 ± 8 151 ± 27 161 ± 13 136 ± 41 136 ± 3

196 ± 5 112 ± 10 ND ND ND ND 70 ± 13 ND 76 ± 5

195 ± 9 54 ± 5 71 ± 1 ND ND ND 32 ± 15

204 ± 8 ND ND ND ND ND ND ND

189 ± 3 167 ± 8 105 ± 3 19 ± 2 20 ± 2 24 ± 5 101 ± 19 ND ND

208 ± 6 89 ± 2 ND 22 ± 0 ND 35 ± 5 ND ND ND

191 ± 5 106 ± 5 ND ND 40 ± 6 20 ± 2 130 ± 25 ND ND

203 ± 6 201 ± 14 186 ± 5 170 ± 1 185 ± 24 176 ± 9 182 ± 36 196 ± 7 195 ± 13

180 ± 15 ND ND ND ND ND ND ND ND

188 ± 5 ND ND ND ND ND ND ND ND

170 ± 7 130 ± 8 ND 67 ± 7 98 ± 4 129 ± 7 105 ± 18 ND ND

195 ± 7 190 ± 13 ND 183 ± 1 180 ± 18 152 ± 12 179 ± 39 175 ± 9 194 ± 8

183 ± 7 182 ± 12 191 ± 4 71 ± 6 168 ± 13 153 ± 11 157 ± 30 176 ± 8 134 ± 11

percent, although in one case the authors used a similar concentration (15.6 mg/L) as in our research (20 mg/L). To use the toxicity test, we diluted the samples from the two media in different ways. Theoretical (initial) concentrations of individual CBAs in the reaction mixture for MEG i

Samples of LNNM media were 0.5 and 0.25 ␮g/ml, respectively. Since we only detected a reduction in toxicity in the initial trials, the dilution of the samples was set to approx. 90% inhibition for controls. Evaluation of the acute toxicity test with V. fischeri was performed by comparison of sample inhibition

Table 2 Restmøngger af CBA i varmedræbte kontroller og after incubation after de tested svampestammer i MEG-medijer (ND: ikke påvist). Mængde CBA (␮g pr. flaske) MEG – 7 dana Kontrola B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor MEG – 14 dana Kontrola B. adusta D. squalens lacteus P chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor MEG – 21 dan Kontrol B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T.si

2-CBA

3-CBA

4-CBA

2,3-CBA

2,4-CBA

2,5-CBA

2,6-CBA

3,4-CBA

3,5-CBA

2,3,5-CBA

2,3,6-CBA

2,4,6-CBA

211 ± 4 201 ± 20 ND ND ND 71 ± 9 ND 90 ± 22

195 ± 4 138 ± 64 97 ± 6 ND ND 44 ± 4 ND ND ND

209 ± 2 162 ± 69 ND ND ND ND ND ND ND

197 ± 2 200 ± 8 ND ND 52 ± 5 56 ± 7 127 ± 7 ND 42 ± 2

207 ± 6 192 ± 25 131 ± 11 46 ± 2 ND ND 66 ± 5 ND 89 ± 2

195 ± 6 189 ± 17 ND 70 ± 27 ND 47 ± 10 145 ± 9 ND 39 ± 1

206 ± 4 200 ± 9 205 ± 20 204 ± 4 209 ± 7 213 ± 7 214 ± 14 201 ± 12 202 ± 7

195 ± 5 154 ± 39 ND ND ND ND ND ND ND

199 ± 4 175 ± 32 ND ND ND ND ND ND ND

175 ± 2 166 ± 13 ND 86 ± 5 ND 76 ± 6 132 ± 10 ND ND

172 ± 6 177 ± 8 151 ± 1 151 ± 4 167 ± 8 172 ± 7 156 ± 11 167 ± 18 154 ± 2

193 ± 5 193 ± 7 160 ± 13 114 ± 11 185 ± 6 188 ± 5 150 ± 14 190 ± 15 174 ± 3

210 ± 30 208 ± 3 ND ND ND ND 37 ± 9 ND 84 ± 16

186 ± 32 191 ± 2 ND ND ND 34 ± 1 ND ND ND

197 ± 24 211 ± 1 213 ± 3 NA NA NA NA NA NA

199 ± 27 162 ± 51 101 ± 6 ND ND 40 ± 8 91 ± 19 ND ND

206 ± 26 203 ± 7 124 ± 6 ND ND 39 ± 5 58 ± 3 ND ND

195 ± 21 194 ± 1 ND ND ND 60 ± 15 117 ± 19 ND ND

212 ± 25 202 ± 12 ND 98 ± 10 226 ± 4 221 ± 2 231 ± 11 188 ± 17 197 ± 5

176 ± 26 175 ± 3 ND ND ND ND ND ND ND

181 ± 25 186 ± 1 ND ND ND ND ND ND ND

164 ± 28 151 ± 10 ND ND ND 86 ± 17 115 ± 15 ND ND

178 ± 26 148 ± 28 145 ± 3 107 ± 12 161 ± 5 162 ± 4 170 ± 7 151 ± 25 150 ± 7

193 ± 24 192 ± 1 166 ± 1 ND 181 ± 2 189 ± 11 150 ± 4 170 ± 23 169 ± 6

209 ± 15 140 ± 53 ND ND ND ND 22 ± 1 ND 121 ± 30

175 ± 17 47 ± 2 ND ND ND ND ND ND ND

188 ± 11 42 ± 9 197 ± 5 ND 93 ± 19 29 ± 8 ND ND 136 ± 22

186 ± 13 173 ± 26 ND ND ND 55 ± 9 86 ± 6 ND ND

193 ± 9 133 ± 46 74 ± 14 ND ND 46 ± 1 61 ± 4 ND ND

182 ± 8 138 ± 54 54 ± 9 ND ND 52 ± 3 126 ± 14 ND ND

213 ± 9 200 ± 6 ND 88 ± 20 194 ± 23 236 ± 3 200 ± 13 187 ± 27 206 ± 11

172 ± 10 80 ± 13 ND ND 59 ± 5 56 ± 0 ND ND ND

177 ± 9 105 ± 10 ND ND 75 ± 4 ND ND ND ND

157 ± 6 137 ± 15 ND ND ND 82 ± 1 119 ± 12 ND ND

166 ± 12 170 ± 4 140 ± 2 96 ± 9 142 ± 19 174 ± 3 187 ± 23 150 ± 25 160 ± 5

181 ± 8 183 ± 7 151 ± 14 ND 167 ± 14 195 ± 2 171 ± 18= 162 ± 34 172 ± 6

390

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Table 3. Inhibition of luminescence in heat-killed controls and after incubation of tested fungal strains in MEG and LNNM media after 21 days of incubation. Luminescence inhibition (%)

LNMM

Kontrol B. adusta D. squalens I. lacteus P. chrysosporium P. magnoliae P. ostreatus P. cinnabarinus T. versicolor

89,4 72,1 11,9 27,2 44,9 41,9 27,5 11,3 14,5

± ± ± ± ± ± ± ± ± ± ±

MEG 8,2 4,3 8,5 13,9 15,4 10,1 8,3 2,7 4,0

94,5 85,9 35,0 7,4 38,2 52,8 34 ,8 20,8 36,3

GRUPA3 GRUPA3

COOH ± ± ± ± ± ± ± ± ± ± ±

1,0 12,5 9,4 3,0 30,7 17,7 2,8 5,6 7,5

Clx COOH

Clx

CHO HO luminescence with their respective controls (Table 3). As mentioned above, the toxicity test revealed that the tested fungi were generally able to reduce the measured acute toxicity, suggesting that the CBA degradation products either did not accumulate or were less toxic than the original CBA.

Clx COOCH3

CH3O

Clx

I GIVE

CH2OH 3.2. Detection of CBA degradation products CBA metabolites were analyzed by GC-MS, and their structures were proposed by comparison of mass spectra with data in the NIST 08 library and independent interpretation of the fragmentation pattern. Furthermore, unknown metabolite structures were investigated using MS/MS (product scanning) to elucidate the fragment sequence. The mass spectral characteristics of the detected CBA degradation products are listed in Table 4. Some of the metabolites were detected after trimethylsilylation (eg chlorobenzyl alcohols), and several of them were confirmed by comparison with available chemical standards. All intermediates were detected only in trace amounts, suggesting that none of them accumulated during degradation. The group of detected intermediates includes chlorobenzaldehydes, chlorobenzyl alcohols, methyl esters of chlorobenzoic acid and methoxy or hydroxy derivatives of these structures. Metabolites were found in different cultures of fungal strains, where representatives of monochloro- and dichlorobenzaldehyde and alcohol, as well as methyl esters of di-CBA, were found in all cultures. The methyl ester representative of tri-CBA was detected only in the culture of I. lacteus, but the methoxy derivatives of tri-CBA and di-CBA methyl esters were found in all fungi. Trichlorinated hydroxybenzyl alcohols were also detected in all fungal cultures. A possible scheme of the fungal CBA degradation pathway constructed from the detected metabolites is shown in the figure. 1. The results are generally similar to those of Kamei et al. [14], who studied the transformation of 4,4-dichlorodiphenyl by Phanerochaete sp. MZ 142, where these authors demonstrated the formation of 4-CBA, methyl ester of 4-CBA and further reduced transformation products: 4-chlorobenzyl alcohol and 4-chlorobenzaldehyde. Such a reduction mechanism could be explained by the action of the intracellular aryl alcohol oxidase system [30]. Matsuzaki et al. showed that enzymes that could be involved in the transformation, namely aryl alcohol dehydrogenase, aryl aldehyde dehydrogenase and also cytochrome P-450 from P. chrysosporium, were up-regulated after the addition of BA to the fungal culture [31]. Other types of transformation products, namely hydroxyl and methoxy derivatives, found in our study were already described by Matsuzaki and Wariishi after the transformation of BA by P. chrysosporium [18]. Metabolites detected include 4-hydroxy, 2-hydroxy and 4-hydroxy-2-methoxy derivatives. In another work, the authors showed that heterologously expressed cytochromes P-450 from the CYP53 family of P. chrysosporium, Aspergilus niger and Rhodotorula minuta were able to hydroxylate BA at position 4 [32]. Mediated by P-450

CH3O

Clx

CH2OH

Clx

CH2OHHO

Clx

CH3O

Clx

Fig. 1. Proposed pathway of CBA degradation by ligninolytic fungi.

Hydroxylation of BA at other sites in fungi has not been reported to date. In addition, the authors used quantitative PCR to show that cytochrome expression is regulated by the presence of BA at the transcriptional level. Induction of cytochrome P-450 BA and also 3 and 4-CBA has been reported elsewhere [17]. Measurement of extracellular ligninolytic activities of fungi in this study showed that most of the activities were suppressed or the maximum activity peaks were delayed during cultivation due to the presence of CBA. Only rare cases where the situation was different were recorded for T. versicolor MnP and laccase activities, which were significantly induced in MEG and LNNM media, respectively. In particular, laccase activity increased from 20 U/L to 230 U/L. These results indirectly confirm that ligninolytic enzymes may not play a key role in CBA degradation. 3.3. Decomposition of CBA in the soil The soil decomposition experiment was monitored after 30 and 60 days, and the residual concentrations after the application of the fungal strains are shown in Figure 2. The results show that, except for the P. cinnabarinus and T. versicolor strains, which decomposed only 30% and 30% respectively %, or 39% of total CBA within 60 days of incubation, all other tested strains were able to remove CBA from soil in the range of 85-99% of total CBA. The results are partially opposite to the experiments with liquid cultures, as P. cinnabarinus was among the most degrading strains in both liquid media. On the other hand, B. adusta proved effective in soil, while this strain was among the least degrading in liquid media. I. lacteus proved to be the most effective

Table 4. Retention data and electron mass impact spectral characteristics of CBA metabolites. MW to CI

fragment ion m/z (relative intensity)

Structure proposal

5,431 5,494 5,582 7,603 7,724 7,913 8,329 8,532 8,875 9,445 9,685 10,883 11,203 11,729 11,8304 11,8304

140 140 140 174 174 174 174 174 214 214 214 204 204 204 204 204

o-chlorobenzaldehyde m-chlorobenzaldehyde p-chlorobenzaldehyde 3,5-dichlorobenzaldehyde 2,4-dichlorobenzaldehyde 2,5-dichlorobenzaldehyde 2,3-dichlorobenzaldehyde MS dichlorobenzaldehyde 3,5-dichlorobenzaldehyde MS-dichlorobenzaldehyde 3,5-dichlorobenzaldehyde 3,3-dichlorobenzaldehyde -dichlordechlor-benzaldehyde-3,3-dichlordehyd-3,3-aldehyde MS o-chlorobenzylalcohol alcohol 2,6-dichlorobenzoesir-methylester 3,5-dichlorobenzoesir-methylester 2,4-dichlorobenzoesir-methylester 2,5-dichlorobenzoesir-methylester 3 ,4-dichlorobenzoesir-methylester

12.651

204

12,945 13,139 13,27 13,604 14,091

170 248 248 248 238

14.288 15 15.801

248 248 238

16.391

238

18.625

310

18.699

298

18.741

310

19.695

234

19,82 20,46 20,592

186 298 310

20.688 21.057 21.526 22.693 23.481

298 298 234 298 234

27.313 29.207

268 312

142 (23,8), 141 (36,6), 140 (73,7), 139 (99,9), 111 (55,0), 75 (32,0), 51 (19,6), 50 (29,8) 142 (20,6), 141 (314,1), (66,6), 139 (99,9), 113 (18,4), 77 (22,7), 75 (33, 5), 74 (19,1) 142 (16,0), 141 (37,1), 140 (49,4), 139 (99,9), 13 16,8), 111 (49,6), 77 (15,1), 74 (16,8) 176 (62,4), 174 (70,3), 173 (99,9), 145 (47,0), 139 (54,2), 111 ( 50,1), 115), 74 (52,6) 176 (39,5), 175 (70,2), 174 (61,4), 173 (99,9), 147 (16,9), 145 (25,9), 75 (18,0), 74 (15,0) 176 (38). 175 (68,4), 174 (61,3), 173 (99,9), 111 (25,0), 75 (61,0), 74 (45,9), 50 (25,3) 176 ( 37,5), 175 (69,0), 174 (613,8), (99,9), 147 (21,9), 145 (37,4), 75 (36,4), 74 (26, 3) 176 (38,9), 175 (69,5), 174 (64,9), 173 (99,9), 147 (28,1), 14 43,0), 75 (29,1), 74 (24,9) 201 (34,0), 199 (99,9), 179 (18,5), 163 (30,7), 127 (25,4), 125 (82,5), 89 ( 25,1), 73 ( ) 201 (31,8), 199 (90,9), 179 (33,4), 171 (19,9), 169 (60,1), 127 (30,2), 125 (99,9), 89 (32,7) 201 (20,8), 199 (5) 179 (24,2), 169 (20,5), 127 (31,9), 125 (99,9) ) ), 89 (24,2), 73 (12,5) 206 (32,5), 204 (43,4), 177 (9,9), 175 (6134). (100), 147 (9,6), 145 (7,9), 109 (7,8), 75 (33,9) 208 (2,0), 206 (14,1), 204 (20,5) ) ), 177 (8,9), 175 (63,1), 173 (100), 147 (21,1), 145 (33,2), 109 (17,8), 75 (16,0) 208 ( 5,4), 206 (12,9), 204 (20,7), 177 (9,9), 175 (61,1), 173 (173) ), 147 (15,5), 145 (29 , 0), 109 (16,2), 75 (95,6) 208 (3,7), 206 (18,3), 204 (25,7), 177 (11,9), 175 (62,1 ) , 173 (93. , 147 (6,4), 145 (19,3), 109 (17,1), 75 (100) 208 (2,7), 206 (19,8), 204 (32, 4) ), 177 (10,3), 175 (64,6), 173 (10), 147 (19,7), 145 (35,1), 109 (22,9), 74 (27,6) 208 (1,3), 206 (16,3), 204 (19,0), 177 (19,0), 175 (56,9), 173 (1090), (29,4), 147 (47, 7), 145 (19,0), 109 (14,4), 75 (97,4) 172 (19,4), 171 (35,0), 170 (56,3), 169 (100), 141 (6) ,8), 12 13,6), 111 (11,7), 77 (15,5) 235 (67,6), 233 (99,9), 205 (18,5), 203 (25 ,6), 161 (41,4), 159 (64,2), 123 (18,4), 123 (18,4), . ) 235 (58,0), 233 (84,2), 161 (61,1), 159 (99,9), 123 (13,8), 103 (29,1), 73 (12,7) 235 (70,9), 233 (99,9), 253,3 (2). 203 (45,8), 161 (58,5), 159 (89,7), 147 (27,7), 123 (21,1) 242 (6,4), 240 (19,3), 238 ( 19,7), 211 (29,3), 209, 295. (100), 183 (3,4), 181 (14,1), 179 (14,5), 143 (11,0), 109 ( 12,8), 74 (12,9) 235 (69,6), 233 (99,9), 205 (17,8), 20 25,3), 161 (46,3), 159 (71,1) ) ), 123 (17,5), 103 (22,8) 235 (68,5), 233 (94,6), 203 (16,7), 161 (67,3), 159 (95,9) , . ), 73 (28,5), 59 (34,2) 242 (7,1), 240 (25,7), 238 (25,1), 211 (31,5), 209 (98,1), 207 (100), 183 (5,7), 181 (12,6), 179 (14,5), 143 (10,5), 109 (11,7), 74 (14,1) 242 (9, 3), 240 (29,4), 238 (29,5), 211 (29,7), 209 (95,4), 207 (10) 183 (6,8), 181 (21,6), 179 (20,7), 143 (14,2), 109 (15,6), 74 (14,5) 271 (32,4), 269 (100), 267 (76,8), 241 (15,8) ) ), 241 (15,8), (45,5), 237 (42,8), 197 (13,7), 195 (46,6), 193 (44,2), 157 (14,2) , 125 (5,5), 123 (15,8), 93 (15,9) 285 (33,2), 28 100), 281 (99,5), 239 (49,1), 237 (50, 6) ), 209 (80,6), 207 (83,3), 205 (2,8), 165 (13,3), 167 (33,3) 271 (32,8), 26), 267 ( 87, 2), 241 (13,0), 239 (38,3), 237 (37,2), 197 (22,0), 195 (68,5), 193 (65,9), 157 (14 ,6) ), 125 (5,6), 123 (1. ), 93 (28,1) 238 (4,3), 236 (41,1), 234 (67,2), 207 (15,4 ), 205 (84,6), 203 (100), 162 (14,1), 160 (18,6), 111 (15, 97 (31,9) 188 (7,2), 186 (25,8) ), 157 (31,6), 155 (100), 127 (17,8), 99 (13,7) 285 (31,5), 283 (100), 281 (9290).(73,8), 237 (36,1), 209 (68,2), 207 (68,5), 167 (12,6), 165 (33,8) 271 (35,5), 269 (100), 267 (99, 1), 241 (10,0), 23 31,1), 237 (31,1), 197 (26,8), 195 (81,5), 193 (82,1), 157 (16,7) , 125 (6,4), 123 (18,0), 93 (39,2) 285 (33 (31,6), 285 (31,6) ), 281 (95,3), 239 (73,8 ), 237 (34,6), 209 (47,7), 207 (43,3), 167 (14,8), 165 (33,0) 285 (34,4), 283 (100), 25) (9. , 239 (36,8), 237 (32,4), 209 (51,5), 207 (54,4), 167 (13,2), 165 (35,6) 238 (3,0), 236 (22,3), 234 (30,4), 207 (10). 205 (68,4), 203 (100), 162 (5,3), 160 (8,3), 111 (11,2), 97 (17,0) 285 (26,1), 283 (100) , 281 (90,9), 239 (52,1), 2 (49,7), 209 (74,2), 207 (64,2), 167 (19,7), 165 (48,1) 238 (5,1), 236 (21,3), 234 (29,5), 207 (11,5), 205 (70,5), 20 100), 162 (10,9), 160 (11,9 ) ), 111 (15,4), 97 (19,9) 270 (15,1), 268 (15,3), 241 (28,3), 239 (95,8), 237 (100) 31) , 314 (13,6), 301 (42,6), 299 (100), 297 (97,3), 271 (24,3), 269 (76,7), 267 (73,4), 227 ( 31), 8), 225 (75,6), 223 ( 88,5)

Type of production

Trimethylsilylering Trimethylsilylering Trimethylsilylering Trimethylsilylering

2,3-dichlorobenzylalcohol methylester α-chloro-α-methoxybenzaldehyde TMS 2,5-dichlorobenzylalcohol TMS 2,4-dichlorobenzylalcohol TMS 3,5-dichlorobenzylalkohol 2,4,6-trichlorobenzylalkohol m

Trimethylsilylering Trimethylsilylering Trimethylsilylering Trimethylsilylering

TMS 2,3-dichlorobenzyl alcohol TMS 3,4-dichlorobenzyl alcohol 2,3,6-trichlorobenzoic acid methyl ester

Trimethylsilylering Trimethylsilylering Trimethylsilylering

2,3,5-trichlorobenzoesir methyl ester TMS α,β,β-triclorbenzil alcohol

Trimethylsilylation

TMS α,α,β-trichloro-α-hydroxybenzilalkohol

Trimethylsilylation

TMS α,β,β-triklorbenzilalkohol

Trimethylsilylation

Mr. Musician i sur. / Journal of Hazardous Materials 196 (2011) 386–394

tR (minute)

α,β-diklor-α-methoxybenzoesiremetilester α-klor-α-hydroxybenzoesiremetilester TMS α,α,α-triklor-α-hydroxybenzilalkohol TMS α,β,β-triklorbenzilalkohol

Methylering Trimethylsilylering Trimethylsilylering

TMS α,α,α-triklor-α-hydroxybenzilalkohol TMS α,α,β-triklor-α-hydroxybenzilalkohol metil-α,β-diklor-α-methoxybenzoate TMSα,α,β-triklor-α-hydroxybenzilmetilester α,β af β-dichlor-α-methoxybenzoesir

Trimetilsililiranje Trimethylsilylering

α,α,β-triklor-α-methoxybenzoesiremethylester TMS α,α,β-triklor-α-methoxybenzilalkohol

Trimethylsilylation

391

392

Mr. Musician i sur. / Journal of Hazardous Materials 196 (2011) 386–394

Fig. 2. Residual concentrations of CBA in contaminated soil after incubation with tested fungal strains: A – 30 days. B – 60 days.

soybeans in soil where this fungus had already depleted 98% of CBA within 30 days. A possible explanation for the discrepancy between these results and the model conditions in the wet medium and the soil degradation experiment lies in the different abilities of fungi to penetrate the contaminated soil [33]. Therefore, we attempted to estimate the relative amount of fungal biomass using ergosterol analysis in soil samples with CBAs and also without the addition of contaminants (Figure 3). Despite the high variability of the data, the results show that the fungi that showed the greatest reduction in CBA (I. lacteus, P. ostreatus and B. adusta) were also strains capable of significantly colonizing contaminated soil. The only exception is P. magnolia, where we discovered one

Significantly lower amount of ergosterol despite high CBA removal (99% within 60 days). The parameters of the Flash Kinetic Toxicity Test were adjusted to also include the detection of potential increases in toxicity. The data obtained by this test are shown in Fig. 4. The best results in terms of inhibition reduction were obtained with I. lacteus and P. ostreatus strains, which corresponds to their CBA degradation efficiency. On the other hand, unexpected results were observed with B. adusta and P. magnolia strains where, despite their high degradation rate, the detected residual toxicity was not significantly different from the controls (t-test, P <0.05). These results of the Flash test are consistent with the results of the toxicity assessment i

Fig. 3. Concentrations of ergosterol in uncontaminated soil and in soil contaminated with CBA after 30 and 60 days of incubation.

Mr. Musician i sur. / Journal of Hazardous Materials 196 (2011) 386–394

393

Fig. 4. Rapid analysis of luminescence inhibition in contaminated soil (control) and in soil with tested fungal strains after 60 days of incubation.

liquid cultures, where residual toxicity was also detected in these fungal cultures. This could be explained by the formation and accumulation of toxic metabolites, and it is probably for the same reason that significantly increased toxicity was observed for T. versicolor. 4. Conclusion To our knowledge, this is the first paper that provides a general description of the ability of ligninolytic fungi to biodegrade CBA, which are critical toxic and highly resistant metabolites in the pathways of bacterial biodegradation of polychlorinated biphenyls. The ability of the sponges was tested in wet conditions and also checked in contaminated soil. The tested fungal strains were capable of degrading CBA in the soil in the range of 85-99% within 60 days, with I. lacteus proving to be the most efficient degrading strain in both tested conditions. Several new degradation products were identified when mainly methoxy and hydroxyl derivatives were produced together with the reduced forms of the starting acids. The results indicate that fungi are likely capable of transforming CBA through multiple pathways with significant reductions in toxicity during the process. The promising degradation results from this study highlight the need for further research, particularly to determine the involvement of different enzymatic mechanisms, to improve the understanding of degradation mechanisms. The results for the liquid medium and the subsequent soil test indicate that the presence of the bioremedial organism is critical. In soil, i.e. in conditions of limited pollutant bioavailability, active soil colonization is equally important. Acknowledgments This work was supported by grants no. 2B06156 of the Ministry of Education, Youth and Sports of the Czech Republic and no. 525/09/1058 from the Czech Science Foundation and institutional research concept no. AV0Z50200510. Literature [1] J.A. Field, R.S. Alvarez, Microbial transformation and degradation of polychlorinated biphenyls, Environ. Pollution. 155 (2008) 1-12.

[2] S.R. Sørensen, M.S. Holtze, A. Simonsen, J. Aamand, Degradation and mineralization of nanomolecular concentrations of the herbicide dichlorovenil and its persistent metabolite 2,6-dichlorobenzamide by Aminobacter spp. isolated from soil treated with dichlorovenil, Appl. Surround. Microb. 73 (2007) 399-406. [3] T. Gichner, P. Lovecká, B. Vrchotová, Genomic damage induced in tobacco plants by chlorobenzoic acids—metabolic products of polychlorinated biphenyls, Mutat. Res. 657 (2008) 140-145. [4] M. Muccini, A.C. Layton, G.S. Sayler, T.W. Schultz, Aquatic toxicities of halogenated benzoic acids to Tetrahymena pyriformis, B. Environ. Contam. X. 62 (1999) 616-622. [5] Y.X. Zhao, G.D. Ji, M.T.D. Cronin, J.C. Dearden, A QSAR study of the toxicity of benzoic acid to Vibrio scheri, Daphnia magna and carp, Sci. The whole environment. 216 (1998) 205-215. [6] P.Y. Lee, C.Y. Chen, Toxicity and quantitative structure-activity relationships of benzoic acid to Pseudokirchneriella subcapitata, J. Hazard. Mater. 165 (2009) 156-161. [7] K. Svobodová, M. Plaˇcková, V. Novotná, T. Cajthaml, Estrogenic and androgenic effects of PCBs, their chlorinated metabolites and other endocrine disruptors assessed using two in vitro yeast studies, Sci. In total. Surround. 407 (2009) 5921-5925. [8] D.H. Pieper, M. Seeger, Bacterial metabolism of polychlorinated biphenyls, J. Mol. Microbial. Biotechnology. 15 (2008) 121-138. [9] S.A. Adebusoye, F.W. Picardal, M.O. Ilori, O.O. Amund, Effect of chlorobenzoic acid on the growth and degradation potential of PCB-degrading microorganisms, World J. Microb. Biot. 24 (2008) 1203-1208. [10] K. Furukawa, H. Fujihara, Microbial degradation of polychlorinated biphenyls: biochemical and molecular features, J. Biosci. It was a meadow. 105 (2008) 433-449. [11] D.P. Barr, S.D. Aust, Mechanisms used by white rot fungi to degrade pollutants, Environ. Sci. Technol. 28 (1994) 78A-87A. [12] C. Pinedo-Rivilla, J. Aleu, I.G. Collado, Biodegradation of pollutants by fungi, Curr. Org. Chem. 13 (2009) 1194-1214. [13] N. Kasai, S. Ikushiro, R. Shinkyo, K. Yasuda, S. Hirosue, Metabolism of monoand dichloro-dibenzo-p-dioxins by Phanerochaete chrysosporium cytochromes P450, Appl. Microbiol. Biot. 86 (2010) 773-780. [14] L. Kamei, R. Kogura, R. Kondo, Metabolism of 4,4-dichlorobiphenyl by white rot fungi Phanerochaete chrysosporium and Phanerochaete sp. MZ142, acc. Microbiol. Biot. 72 (2006) 566-575. [15] T. Cajthaml, Z. Kˇresinová, K. Svobodová, M. Möder, Biodegradation of endocrine disrupting compounds and suppression of estrogenic activity by ligninolytic fungi, Chemosphere 7 (2009) 745-750. ˇ sec, M. Möder, Degradation products on the metabolic [16] T. Cajthaml, P. Erbanová, V. Saˇ path of degradation of benz[a]anthracene by a ligninolytic fungus, Chemosphere 64 (2006) 560–564. [17] D. Ning, H. Wang, Y. Zhuang, Induction of functional cytochrome P450 and its involvement in benzoic acid degradation by Phanerochaete chrysosporium, Biodegradation 21 (2010) 297-308. [18] F. Matsuzaki, H. Wariishi, Functional diversity of cytochrome P450 of the white rot fungus Phanerochaete chrysosporium, Biochem. Bioph. Res. co. 324 (2004) 387-393. ˇ Novotny, ˇ sec, Degradation of PAHs ´ V. Saˇ [19] T. Cajthaml, P. Erbanová, A. Kollmann, C. by ligninolytic enzymes of Irpex lacteus, Folia Microbiol. 53 (2008) 289-294. ˇ sec, M. Matucha, Degradation of anthracene [20] B.R.M. Vyas, S. Bakowski, V. Saˇ selected white rot fungi, FEMS Microbiol. Ecol. 14 (1994) 65-70. [21] E. De Jong, A.E. Cazemier, J.A. Field, J.A. De Bont, Physiological role of chlorinated aryl alcohols de novo biosynthesized by the white rot fungus Bjerkandera sp. strain BOS55, Appl. Surround. Microbiol. 60 (1994) 271-277.

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Journal of Hazardous Materials 196 (2011) 395-401

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Comparison of o-Toluidine Degradation by Fenton, Electro-Fenton and Photo-Electro-Fenton Processes Jin Anotai a, Somporn Singhadech b, Chia-Chi Suc, Ming-Chun Lu c,* National Center of Excellence for Environmental and Hazardous Waste Management, Department of Engineering of Environment, Faculty of Engineering, King Mongkut University of Technology Thonburi, Bangkok 10140, Thailand b Department of International Graduate Programs in Environmental Management (Hazardous Waste Management), Chulalongkorn University, Bangkok 10330, Thailand c Department of Resource Management Envi -Nan University of Pharmacy and Science, Tainan 717, Taiwan

article

i n f o

Article history: Received May 27, 2011 Received in revised form September 5, 2011 Accepted September 11, 2011 Available online September 22, 2011 Keywords: Box–Behnken electro-Fenton scheme Hydroxyl radicals Photoelectro-Fenton o-Toluide process

a b s t r a c t Statistical Box-Behnken experimental design (BBD) was used to investigate the degradation of o-toluidine by the electro-Fenton process. This method can be used to determine the optimal conditions in multivariate systems. Fe 2+ concentration (0.2-1.0 mM), H 2 O 2 concentration (1-5 mM), pH (2-4) and current strength (1-4 A) were selected as independent variables. The removal efficiency of o-toluidine and chemical oxygen consumption (COD) is represented by the response function. The result with 2-level factorial design shows that pH and concentrations of Fe2+ and H2O2 were the main parameters. Among the main parameters, the removal efficiency of o-toluidine and COD was significantly affected by pH and Fe2+ concentration. From Box-Behnken design predictions, the optimum conditions in the electro-Fenton process for the removal of 90.8% o-toluidine and 40.9% COD were found to be 1 mM Fe2+ and 4.85 mM H2O2 at pH 2. the optimum conditions, experimental data showed that the removal efficiency of o-toluidine and COD in the electro-Fenton process and the photo-electro-Fenton process was over 91% and 43%, respectively, after 60 minutes of reaction. The removal efficiency of o-toluidine and COD in the Fenton process is 56% and 27%, respectively. © 2011 Elsevier B.V. All rights reserved.

1. Introduction O-toluidine is an important aromatic amine used in the paint and rubber industries. However, short-term exposure to otoluidine can cause methemoglobinemia, while long-term or repeated exposure to o-toluidine can be carcinogenic to humans [1]. can cause bladder cancer [2]. It is difficult to completely purify wastewater containing o-toluidine due to its resistance to biodegradation. Currently, advanced oxidation processes (AOP) are used for wastewater treatment, especially in cases where pollutants are difficult to remove by biological or physicochemical processes [3-7]. AOPs are based on the creation of a strong oxidant, the hydroxyl group (• OH), which can react with most organic pollutants and then degrade them [8,9]. The Fenton process is one of the most widely used AOPs due to its low investment costs [10]. The Fenton reaction is shown below: Fe2+ + H2 O2 → • OH + Fe3+ + OH−

∗ Corresponding author. Phone: +886 6 266 4911; fax: +886 6 266 3411. E-mail address:[email protected](M.-C. Lu). 0304-3894/$ – see front page © 2011 Elsevier B.V. All rights reserved. doi:10.1016/j.jhazmat.2011.09.043

(1)

But the iron precipitate produced by the Fenton process requires further processing and disposal, which is a major disadvantage of this process. This major drawback can be solved by coupling the Fenton process with electrical discharge, the so-called "electro-Fenton (EF) process". The advantage of the electrochemical Fenton process is that it produces much less iron precipitate than the traditional Fenton process. In this process, iron ions (Fe3+) are efficiently electrogenerated into iron ions (Fe2+), as shown in Eq. (2); This can be expressed in terms of current performance. Fe3+ + e− → Fe2+

(2)

The capability of the electro-Fenton process was confirmed by Harrington and Pletcher [11], with more than 90% removal of chemical oxygen consumption (COD) with energy efficiency greater than 50% and acceptable energy consumption. The efficiency of the electro-Fenton process can be improved by using UV or visible light in a process known as photo-electro-Fenton (PEF). This improvement is due to a higher rate of • OH production from Fe(OH)2+ photoreduction (equation (3)) and photodecomposition of the complex from Fe3+ reactions (equation (4)) [12-15 ] Fe (OH) 2+ + hv → Fe2+ + • OH R(CO2)-Fe

+ hv →

R(• CO

(3) + Fe(II) →

•R

+ CO2

(4)

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Κανονικά, οι διεργασίες τύπου Fenton [16-18] επηρεάζονται από το pH και τις αρχικές συγκεντρώσεις Fe2+ και H2O2. Για τον προσδιορισμό των βέλτιστων συνθηκών για την αποικοδόμηση της ο-τολουιδίνης και την επίδραση των μεταβλητών στη διαδικασία ηλεκτρο-Fenton, χρησιμοποιήθηκε ο σχεδιασμός Box-Behnken (BBD) σε αυτήν την έρευνα. Το BBD μπορεί να χρησιμοποιηθεί για την εύρεση των βέλτιστων συνθηκών σε πολυμεταβλητά συστήματα [19]. Ο στατιστικός σχεδιασμός ενός πειράματος μειώνει τον αριθμό των πειραμάτων που πρέπει να εκτελεστούν και τον αντίστοιχο χρόνο που δαπανάται και μπορεί να χρησιμοποιηθεί για τη βελτιστοποίηση των παραμέτρων λειτουργίας σε πολυμεταβλητά συστήματα. Λίγες μελέτες έχουν χρησιμοποιήσει το BBD για την αποικοδόμηση αζωχρωστικών και οργανικών ρύπων με τη διαδικασία photo-Fenton [19,20]. Ωστόσο, δεν έχουν δημοσιευθεί μελέτες σχετικά με τη χρήση του BBD για την αποικοδόμηση της ο-τολουιδίνης με τη διαδικασία electro-Fenton. Σε αυτή τη μελέτη, οι βέλτιστες συνθήκες για την αποικοδόμηση της ο-τολουιδίνης και η επίδραση τεσσάρων μεταβλητών (pH, συγκέντρωση Fe2+, συγκέντρωση H2O2 και ρεύμα) στη διεργασία ηλεκτρονίου-Fenton διερευνήθηκαν χρησιμοποιώντας BBD. Καθώς μόνο αρκετοί σημαντικοί παράγοντες εμπλέκονταν στη βελτιστοποίηση, εφαρμόστηκε η μεθοδολογία επιφάνειας απόκρισης (RSM). Επιπλέον, συγκρίθηκαν επίσης οι επιδόσεις αποικοδόμησης ο-τολουιδίνης των συνηθισμένων διεργασιών Fenton, electroFenton και photoelectro-Fenton. 2. Υλικά και μέθοδοι 2.1. Το υλικό και ο αντιδραστήρας ο-τολουιδίνη (99,5%, Merck), υπεροξείδιο του υδρογόνου (H2O2, 35%, Merck) και επτα-ένυδρος θειικός σίδηρος (FeSO4 ·7H2O, Merck) ήταν βαθμού αντιδραστηρίου και χρησιμοποιήθηκαν χωρίς περαιτέρω καθαρισμό. Το Σχ. 1 δείχνει τα τρία είδη αντιδραστήρων. Ο αντιδραστήρας Fenton ήταν κυλινδρικός ανοξείδωτος χάλυβας (διάμετρος: 13 cm, ύψος: 35 cm). Ο συνολικός όγκος του αντιδραστήρα ήταν 3,5 L. Ο αντιδραστήρας electro-Fenton, ένας κυλινδρικός αντιδραστήρας, λειτουργούσε σε λειτουργία σταθερού ρεύματος. Η άνοδος ήταν επικαλυμμένο με δίχτυ τιτανίου με RuO2/IrO2 (DSA) και η κάθοδος ήταν από ανοξείδωτο χάλυβα. Η άνοδος DSA με εσωτερική διάμετρο 7 cm και ύψος 35 cm, και η κάθοδος είχε εσωτερική διάμετρο 2 cm και ύψος 35 cm. Τα ηλεκτρόδια συνδέθηκαν με συνεχές ρεύμα (DC). Στον αντιδραστήρα φωτοηλεκτρο-Fenton, ένα σύνολο 6 λαμπτήρων υπεριώδους ακτινοβολίας στερεωμένες μέσα σε έναν κυλινδρικό σωλήνα Pyrex (επιτρέποντας σε μήκη κύματος > 320 nm να διεισδύσουν) χρησιμοποιήθηκαν ως πηγή ακτινοβολίας. Οι λάμπες UV συνδέθηκαν στο τροφοδοτικό E-safe, 2003, Switching Power Supply (Max. 300 W), Μοντέλο: LC-B300AT. 2.2. Μέθοδος ανάλυσης Στο πείραμα φωτοηλεκτρο-Fenton, παρασκευάστηκαν συνθετικά λύματα που περιείχαν 1 mM ο-τολουιδίνης και στη συνέχεια το αρχικό ρΗ ρυθμίστηκε με υπερχλωρικό οξύ (HClO4). Μετά τη ρύθμιση του pH, μια προκαθορισμένη ποσότητα καταλυτικού θειικού σιδήρου προστέθηκε στο διάλυμα και στη συνέχεια άναψαν οι λυχνίες UV. Προστέθηκε επίσης Η2Ο2 στον ίδιο χρόνο για να ξεκινήσει η αντίδραση. Επιπλέον, στο πείραμα electro-Fenton, παρασκευάστηκε διάλυμα με 1 mM ο-τολουιδίνης και στη συνέχεια προστέθηκαν ιόντα σιδήρου αφού το pH ρυθμίστηκε στην επιθυμητή τιμή. Εν τω μεταξύ, η παροχή ρεύματος ενεργοποιήθηκε και προστέθηκε υπεροξείδιο του υδρογόνου για να ξεκινήσει η αντίδραση. Δείγματα (1 mL) ελήφθησαν σε προκαθορισμένα χρονικά διαστήματα και εγχύθηκαν αμέσως σε σωλήνα που περιείχε διάλυμα υδροξειδίου του νατρίου για να σβήσει η αντίδραση Fenton αυξάνοντας το pH στο 11. Το δείγμα στη συνέχεια διηθήθηκε (φίλτρο 0,45 ␮m) για να απομακρυνθούν τα ιζήματα και διατηρήθηκε για 12 ώρες πριν από την ανάλυση COD. Αυτή η διαδικασία χρησιμοποιήθηκε για να αποφευχθεί η ποσοτικοποίηση της επίδρασης της συγκέντρωσης H2O2 στην τιμή COD. Το COD αναλύθηκε με μια κλειστή τιτρομετρική μέθοδο παλινδρόμησης με βάση τις τυπικές μεθόδους [21]. Η συγκέντρωση Fe2+ προσδιορίστηκε χρησιμοποιώντας τη μέθοδο 1,10-φαινανθρολίνης [22]. Ο συνολικός οργανικός άνθρακας μετρήθηκε με αναλυτή υγρού TOC Elementar. Η συγκέντρωση της ο-τολουιδίνης

προσδιορίστηκε χρησιμοποιώντας υγρή χρωματογραφία υψηλής απόδοσης (HPLC) με αντλία συστήματος Spectra μοντέλο SN4000 και στήλη Asahipak ODP-506D (150 mm × 6 mm × 5 ␮m). Το όριο ανίχνευσης της ο-τολουιδίνης ήταν 0,005 mM ή 0,535 ppm. Τα οργανικά οξέα αναλύθηκαν χρησιμοποιώντας χρωματογραφία ιόντων Dionex DX-120 με στήλη ανιόντων Ion Pac AS11 στους 30◦ C. 2.3. Το πειραματικό σχέδιο BBD είναι μια κατηγορία περιστρεφόμενων ή σχεδόν περιστρεφόμενων σχεδίων δεύτερης τάξης που βασίζονται σε ελλιπή παραγοντικά σχέδια τριών επιπέδων. Μεταξύ όλων των σχεδίων RSM, το BBD απαιτεί λιγότερες εκτελέσεις [23]. Η έκδοση λογισμικού DesignExpert έκδοση 7.0 (Stat-Ease, Inc., Minneapolis, USA) χρησιμοποιήθηκε για την εύρεση των βέλτιστων συνθηκών αποικοδόμησης της ο-τολουιδίνης με τη διαδικασία electro-Fenton. Οι επιδράσεις των σημαντικών παραγόντων προσδιορίστηκαν από το BBD. Οι σημαντικοί παράγοντες και τα κατάλληλα εύρη που μελετήθηκαν ήταν pH: 2–4, συγκέντρωση Fe2+: 0,2–1,0 mM και συγκέντρωση H2O2: 1–5 mM. Η συγκέντρωση της ο-τολουιδίνης σταθεροποιήθηκε στο 1 mM για όλα τα πειράματα. 3. Αποτελέσματα και συζήτηση 3.1. Επίδραση διαφόρων παραμέτρων στην αποτελεσματικότητα απομάκρυνσης της ο-τολουιδίνης Η συγκέντρωση Fe2+, η συγκέντρωση H2O2, το pH και το ρεύμα επιλέχθηκαν ως πειραματικές συνθήκες για το BBD. Οι αποτελεσματικότητες αφαίρεσης για την ο-τολουιδίνη και το COD αντιπροσωπεύτηκαν από μια συνάρτηση απόκρισης. Ο Πίνακας 1 δείχνει τα δύο επίπεδα των τεσσάρων παραγόντων στο BBD. Οι τιμές των μεταβλητών, τα πειραματικά δεδομένα και τα αποτελέσματα παρουσιάζονται στον Πίνακα 2. Ο μέγιστος ρυθμός απομάκρυνσης της ο-τολουιδίνης ήταν 100% και ο ελάχιστος ήταν 23% (Πίνακας 2). Όταν εφαρμόστηκαν 1,0 mM Fe2+, 5,0 mM Η2Ο2 και 1,0 Α ρεύματος σε ρΗ 2, η απομάκρυνση της ωτλουιδίνης ήταν 94,4% (πείραμα 2). Ωστόσο, καθώς το ρεύμα αυξανόταν από 1,0 Α σε 4,0 Α, η απομάκρυνση της ο-τολουιδίνης αυξήθηκε ελαφρά στο 100% (εκτέλεση 6). Διαπιστώθηκε ότι η ποσότητα του ρεύματος δεν ήταν ευαίσθητη στην εφαρμοζόμενη περιοχή και επομένως μπορούσε να παραμεληθεί. Η συσχέτιση της αποτελεσματικότητας αφαίρεσης ο-τολουιδίνης και COD που λαμβάνεται από το BBD φαίνεται στον Πίνακα 1, όπου υψηλότερη συσχέτιση σημαίνει ότι η παράμετρος έχει υψηλότερη επίδραση στην ο-τολουιδίνη και το COD. Η συσχέτιση μπορεί να είναι τόσο υψηλή όσο 1 ή χαμηλή έως -1. Το αποτέλεσμα δείχνει ότι το ρεύμα έχει μια ελαφρά επίδραση στην ο-τολουιδίνη, με συσχέτιση μόνο 0,098 (Πίνακας 1). Η αποικοδόμηση της ο-τολουιδίνης εξαρτιόταν από την αρχική συγκέντρωση των Fe2+ και H2O2, δείχνοντας υψηλή συσχέτιση στην αποτελεσματικότητα απομάκρυνσης της ο-τολουιδίνης περίπου 0,617 και 0,278 για τη συγκέντρωση Fe2+ και H2O2, αντίστοιχα. Η ίδια τάση παρατηρήθηκε επίσης στην αποτελεσματικότητα αφαίρεσης COD. Αυτή η συσχέτιση δείχνει ότι οι συγκεντρώσεις Fe2+ και H2O2 έχουν θετική επίδραση στις αποτελεσματικότητες αφαίρεσης για ο-τολουιδίνη και COD, υποδεικνύοντας ότι η αύξηση των συγκεντρώσεων Fe2+ και H2O2 αύξησε την αποτελεσματικότητα αφαίρεσης για ο-τολουιδίνη και COD. Το pH έχει αρνητική επίδραση στην απομάκρυνση της ο-τολουιδίνης και του COD, και έτσι οι αποτελεσματικότητες αφαίρεσης για την otoluidine και COD μειώθηκαν με την αύξηση του pH του διαλύματος. Από τις τιμές συσχέτισης, το αρχικό pH και οι συγκεντρώσεις Fe2+ και H2O2 ήταν οι πιο σημαντικοί παράγοντες που επηρέασαν την απομάκρυνση της ο-τολουιδίνης και του COD. Ο Πίνακας 3 δείχνει τα επίπεδα των σημαντικών παραγόντων της αποτελεσματικότητας αφαίρεσης της ο-τολουιδίνης και του COD. Τα αποτελέσματα από το πείραμα αποκάλυψαν ότι η μέγιστη απομάκρυνση της ο-τολουιδίνης ήταν 91,4% και εκείνη της COD ήταν 42% (πείραμα 1) (Πίνακας 4). Οι τιμές συσχέτισης υποδεικνύουν ότι το pH είχε την πιο έντονη επίδραση στην απομάκρυνση της ο-τολουιδίνης και του COD (-0,725 για την ο-τολουιδίνη και −0,593 για το COD) (Πίνακας 3). Η συγκέντρωση Fe2+ είχε συγκρίσιμη επίδραση σε αυτές τις αποκρίσεις και η συγκέντρωση Η2Ο2 είχε μεγαλύτερη επίδραση στην απομάκρυνση του COD από την αποικοδόμηση της ο-τολουιδίνης. Το Σχ. 2 δείχνει το διάγραμμα της επιφάνειας απόκρισης της επίδρασης του pH και της συγκέντρωσης Fe2+ στις αποτελεσματικότητες αφαίρεσης ο-τολουιδίνης και COD. Αυτό

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397

sl. 1. Experimental reactors.

Table 1. Two levels of variables and the correlation value with the removal efficiency of o-toluidine and COD by statistical Box-Behnken design. Variables

Symbol

pH Fe2+ (mM) H2 O2 (mM) Strøm (A)

A B C D

Variable level

link

Lav

high

o-toluidine

COD

2 0,21 1

4 1 5 4

−0,628 0,617 0,278 0,098

−0,517 0,633 0,274 0,259

Table 2. Removal of o-toluidine and COD at two levels of variables in electro-Fenton process with 1 mM o-toluidine designed by BBD. Execution number

Fe2+ (mM)

H2O2 (mM)

Current (A)

Removal of o-toluidine (%)

Cod (%)

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16

4,0 2,0 4,0 4,0 2,0 2,0 2,0 2,0 4,0 4,0 2,0 2,0 4,0 4,0 2,0 4,0

0,2 1,0 0,2 1,0 0,2 1,0 1,0 1,0 1,0 0,2 0,20,20,21,0 0, 2 1, 0

5,0 5,0 5,0 5,0 5,0 5,0 1,0 1,0 1,0 1,0 5,0 1,0 1,0 5,0 1,0 1,0

4,0 1,0 1,0 4,0 1,0 4,0 1,0 4,0 4,0 1,0 4,0 1,0 4,0 1,0 4,0 1,0

17,0 94,4 32,0 56,0 42,0 100 66,4 68,5 45,0 23,0 74,0 40,0 26,5 61,4 56,4 48,0

23,0 57,0 22,0 33,02 8,0 59,5 33,0 46,0 34,0 19,0 38,4 22,5 27,0 36,0 34,6 31,0

The graph shows the negative effect of pH on removal efficiency. The removal of o-toluidine and COD decreased as the pH of the initial solution increased from 2.0 to 4.0 because the oxidation potential of hydroxyl radicals (•OH) and the dissolved fraction of iron species decreased [24,25]. The results also show that increasing the concentration of Fe2+ can improve the removal efficiency of o-toluidine and COD because more Fe2+ reacts with H2O2 and produces more • OH. Analysis of variance (ANOVA) tests were performed for o-toluidine and COD removal to determine suitability

response function and the significance of the effects of independent variables on the response function (table 5). ANOVA shows that the predictability of the model is at the 95% confidence level. "Prob > F" values less than 0.05 indicate a significant effect of the corresponding variable on the response. The result shows that the F values for o-toluidine and COD removal were 11.10 and 8.06, respectively. means that the model is important. There are only 0.15% and 0.43% chances of removing o-toluidine and COD, respectively. that such large F values of the model may be due to noise.

Table 3. Levels of significant factors and correlation value with the removal efficiency of o-toluidine and COD from BBD. An important fact

Symbol

Variable level

link

Lav

Center

high

o-toluidine

COD

pH Fe2+ (mM) H2O2 (mM)

i B C

2 0,21

3 0,6 3

4 1 5

−0,725 0,484 0,294

−0,593 0,455 0,408

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Table 4 Removal of o-toluidine and COD at significant factor levels in electro-Fenton process with 1 mM o-toluidine and 1 A designed by BBD. Execution number

Fe2+ (mM)

H2O2 (mM)

Removal of o-toluidine (%)

Cod (%)

1 2 3 4 5 6 7 8 9 10 11 12 13

2,0 4,0 2,0 2,0 2,0 4,0 3,0 40 3,0 3,0 3,0 4,0 3,0

0,6 0,6 1,0 0,2 0,6 0,6 0,2 0,20,6 1,0 0,2 1,0 1,0

5,0 5,0 3,0 3,0 1,0 1,0 5,0 3,0 3,0 1,0 1,0 3,0 5,0

91,4 31,2 75,0 47,4 63,4 34,0 51,0 15,0 60,0 52,0 42,4 49,4 78,0

42,0 17,6 36,0 16,4 30,0 16,0 24,0 17,0 32,0 21,0 16,4 21,2 36,0

In this case, pH and Fe2+ were significant model terms affecting o-toluidine percentage and COD removal. 3.2. Prediction of the optimal conditions for the degradation of o-toluidine by BBD The aim of this part was to determine the optimal conditions for the maximum removal of o-toluidine and COD by the electro-Fenton process. BBD can provide an empirical relationship between the response function and the variables. The mathematical relationship between the removal of o-toluidine and three important variables can be approximated by a quadratic polynomial equation, and the equations for the removal of o-toluidine and COD by the electro-Fenton process are shown below: o - toluidine removal ( % ) = 64.22 − 11, 02A + 12.36B + 3.25C + 2.73AB − 1.02AC + 8.58BC − 1.17ABC

(5)

COD removal (%) = 30.16 − 2.31A + 7.47B + 0.84C + 0.91AB − 2.81AC + 8.03BC − 1.81ABC

of o-toluidine and COD at each pH value, Fe2+ concentration and H2O2 concentration Based on the coefficients in Eq. (5) and (6) show that the concentration of pH (A) and Fe2+ (B) has negative and positive effects on the removal efficiency of o-toluidine and COD, respectively. In other words, the removal of o-toluidine and COD decreased with pH (A), while it increased with the doses of Fe2+ (B) and H2O2 (C). Dosing Fe2+ had a more profound effect on the removal of o-toluidine and COD compared to H2O2. The Electro-Fenton process uses the electrochemical production of iron ions from iron ions and iron complexes. The iron ions were continuously electrochemically recycled and thus were not depleted during the decomposition of o-toluidine. The concentration of Fe2+ has the greatest effect on the removal of o-toluidine with the highest coefficient (12.36). In this study, the removal of o-toluidine and COD was selected as "maximization", and then the concentrations of Fe2+ and H2O2 and pH were used as "within the range". The software therefore combined these individual measures into an overall desirability function to find the best optimal conditions. The relationship of the final equation between the response function (o-toluidine and COD removal) and the key parameters can be determined from Eq. (7) and (8).

(6)

Fe2+

where A, B and C are respectively pH, concentration and concentration of H2O2. Equations are used to calculate subtraction

o − toluidine removal (%) = 78.74 − 18.45 × pH + 30.81 × [Fe2+] + 3.74 × [H2O2]

(7)

Table 5 ANOVA tests for removal of o-toluidine and COD from BBD. Source

Sum of squares

Inside the blocks

F value

p-value Probability > F

o-Toluidinfjernelse Model A (pH) B (Fe2+ ) C (H2 O2 ) AB AC BC ABC Residualni Cor Total

7790,65 3387,24 3271,84 663,06 1,56 190,44 262,44 14,06 802,07 8592,72

7 1 1 1 1 1 1 1 1 8 15

112,95 3387,24 3271,84 663,06 1,56 190,44 262,44 14,06 100,26

11,10 33,78 32,63 6,61 0,016 1,90 2,62 0,14

0,0015 0,0004 0,0004 0,0330 0,9037 0,2055 0,1443 0,7178

Important Important Important

COD-fjernelse Model A (pH) B (Fe2+ ) C (H2 O2 ) AB AC BC ABC Residualni Cor Total

1808,81 552,25 826,56 155,00 52,56 119,90 68,89 33,64 256,41 2065,22

7 1 1 1 1 1 1 1 1 8 15

258,40 552,25 816,56 155,00 52,56 119,90 68,89 33,64 32,05

8,06 17,23 25,79 4,84 1,64 3,74 2,15 1,05

0,0043 0,0032 0,0010 0,0591 0,2362 0,0891 0,1808 0,3356

Important Important Important

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Fig. 2. 3D representation of the response surface diagram of the effect of pH and Fe 2+ concentration on (a) o-toluidine and (b) COD removal efficiency.

COD removal (%) = 30.41 − 6.58 × pH + 12.63 × [Fe2+ ] + 2.27 × [H2O2]

(8)

According to equations (7) and (8), at the optimal conditions of pH 2, 1 mM Fe2+ and 4.85 mM H2O2, the maximum removal of o-toluidine and COD was 90.8% and 40.9%, respectively. 3.3. Comparison between different processes Optimum conditions were used to test the removal efficiency of o-toluidine and COD in Fenton process, electroFenton process and photoelectro-Fenton process. The results are shown in fig. 3rd fig. Figure 3(a) shows that the otluidine removal efficiency of the three processes was almost the same in the first 2 minutes. After 2 minutes, the Fenton removal of o-toluidine increased slightly. The removal efficiency for o-toluidine was approx. 56% after 60 minutes of reaction. However, the removal efficiency of o-toluidine in the electro-Fenton process and the photo-electro-Fenton process was 91% and 99%, respectively, after 60 minutes of reaction. The removal of o-toluidine was due to the formation of • OH via eq. (1). Additionally, Fe3+ in solution could be regenerated inside the reactor when electrical discharges and UV irradiation were used, allowing some Fe2+ to react with H2O2 to produce • OH. Iron ions are not depleted during the oxidation reaction, as shown in Eq. (2)-(4). Therefore, electro-Fenton process and photo-electro-Fenton process can increase the oxidation rate of o-toluidine. The same result was found for COD removal efficiency, as shown in the figure. 3(b). The COD removal efficiency of the three processes was similar in the first 2 minutes. However, the COD removal efficiency was significantly different between the different processes after two minutes of reaction. COD removal efficiency was 27% for the Fenton process, 45% for the electro-Fenton process and 43% for the photo-electro-Fenton process after 60 minutes of reaction. Fig. Figure 4 shows that maleic and oxalic acids were identified as intermediates of o-toluidine oxidation. Maleic and oxalic acids were found in the electro-Fenton and photoelectro-Fenton processes after 1 minute of reaction, and the concentrations increased with time. Reduction of maleic acid occurred in electro-Fenton and photo-electro-Fenton processes after 45 and 10 minutes, respectively (Figure 4(a)). Oxalic acid reduction occurred in electro-Fenton and photo-electro-Fenton processes after 30 and 45 minutes, respectively (Figure 4(b)). However, after 10 minutes of reaction in the Fenton process, maleic and oxalic acid were found, the concentrations of which increased with time until the end of the reaction. The concentration of maleic and oxalic acid was determined

quickly due to increased • OH concentration and decomposition of o-toluidine. This result shows that the electro-Fenton and photo-electro-Fenton processes have higher efficiency in the degradation of o-toluidine than the traditional Fenton process. The accumulation of intermediates in the photoelectro-Fenton process was lower than that in the electro-Fenton process (Fig. 4(a) and (b)). In addition, the TOC removal efficiencies in the electro-Fenton and photo-electro-Fenton processes were 12% and 31%, respectively (Figure 4(c)). These phenomena show that the intermediate products are effectively mineralized by the action of UV light in the photoelectroFenton process.

Fig. 3. Comparison between different processes for (a) o-toluidine removal and (b) COD removal efficiency. Experimental conditions: 1 mM o-toluidine, 1 mM Fe 2+ and 4.85 mM H 2 O 2 at pH 2. Each cell has two samplings.

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results with 91% removal of o-toluidine and 45% removal of COD in the electro-Fenton process, which indicates the reliability of the method used. Both electric discharge and UV radiation can significantly increase the degradation of o-toluidine. Therefore, at the beginning of the electro-Fenton and photoelectro-Fenton processes, more intermediates such as maleic and oxalic acid were produced than in the traditional Fenton process. Acknowledgments The authors would like to thank the National Science Council of Taiwan for financial support for this research under contract no. NSC 96-2628-E-041-001-MY3. bibliographical references

Fig. 4. Concentrations of (a) maleic acid, (b) oxalic acid and (c) TOC during o-toluidine degradation. Each data has two samples.

4. Conclusions This study investigated the optimization of o-toluidine electro-Fenton treatment using the Box-Behnken experimental design methodology. The results showed that pH and Fe2+ concentrations are significant factors in the removal efficiency of both o-toluidine and COD. The removal efficiency of o-toluidine and COD increases with decreasing pH and increasing Fe2+ concentration. The optimal conditions for maximum removal of o-toluidine and COD (90.8% and 40.9% of predicted, respectively) were 1 mM Fe2+ and 4.85 mM H2O2 at pH 2. It is clear that the calculations of removal of o-toluidine and COD were applicable . the predicted conditions approach the experimental ones

[1] G. Korinth, L. Lüersen, K.H. Schaller, J. Angerer, H. Drexler, Enhancement of transdermal penetration of aniline and o-toluidine in vitro by skin protection creams, Toxicol. In Vitro 22 (3) (2008) 812-818. [2] E. Richter, K. Gaber, U.A. Harréus, C. Matthias, N. Kleinsaser, o-toluidine adducts in human bladder DNA and hemoglobin from the local anesthetic prilocaine, Toxicol. Easy. 164S (2006) S255. [3] J.H. Sunce, S.P. Sun, M.H. A fan, H.Q. Guo, L.P. Qiao, R.X. Sun, Kinetic study of p-nitroaniline degradation by Fenton oxidation process, J. Hazard. Mater. 148 (2007) 172-177. [4] E. Brillas, I. Sirés, M.A. Oturan, Electro-Fenton process and related electrochemical technologies based on Fenton reaction chemistry, Chem. Rev. 109 (2009) 6570–6631. [5] J. Virkutyte, E. Rokhina, V. Jegatheesan, Optimization of electro-Fenton denitrification wastewater model using response surface methodology, Bioresour. Technol. 101 (2010) 1440-1446. [6] S. Mohajeri, H.A. Aziz, M.H. Isa, M.A. Zahed, M.N. Adlan, Statistical optimization of process parameters for landfill leachate treatment using the electro-Fenton technique, J. Hazard. Mater. 176 (2010) 749-758. [7] L.C. Almeida, S. Garcia-Segura, N. Bocchi, E. Brillas, Solar photoelectroFenton degradation of paracetamol using the flow plant with Pt/air diffusion cell connected to a complex parabolic collector: process optimization with response surface methodology, Appl. A cat. B: Environment. 103 (2011) 21-30. [8] W.H. Glaze, J.W. Kang, R.H. Chapin, Chemistry of water treatment processes involving ozone, hydrogen peroxide and ultraviolet radiation, Ozone Sci. Meadow. 9 (4) (1987) 335-352. [9] M.J. Liou, M.L. Lu, Catalytic Degradation of Nitroaromatic Explosives with Fenton's Reagent, J. Mol. A cat. A: Chem. 277 (2007) 155-163. [10] S. Esplugas, J. Gimenez, S. Contreras, E. Pascual, M. Rodriguez, Comparison of different advanced oxidation processes for phenol degradation, Water Res. 36 (4) (2002) 1034-1043. [11] T. Harrington, D. Pletcher, Removal of low levels of organic matter from aqueous solutions using Fe(II) and hydrogen peroxide formed in situ at gas diffusion electrodes, J. Electrochem. Soc. 146 (1999) 2983-2989. [12] H.J. Benkelberg , P. Warneck , Removal of 3-chlorophenol by excitation of a dilute iron(III) solution: evidence for exclusive involvement of Fe(OH)2+ , J. Phys. Chem. 99 (1995) 5214-5221. [13] R. Andreozzi, V. Caprio, R. Marotta, Photooxidation of 2-aminophenol in aqueous solution induced by iron (III) oxide (hyd.): a kinetic study, Water Res. 37 (2003) 3682-3688. [14] M. Fukushima, K. Tatsumi, K. Morimoto, Fate of aniline after photo-Fenton reaction in an aqueous system containing iron(III), humic acid and hydrogen peroxide, Environ. Sci. Technol. 34 (2000) 2006-2013. [15] C.L. Hsueh, Y.H. Huang, C.Y. Chen, A novel iron oxide composite supported on activated alumina as a heterogeneous catalyst for the photooxidative degradation of reactive carbon black 5, J. Hazard. Mater. B 129 (2006) 228-233. [16] H. Zhang, C. Fei, D. Zhang, F. Tang, Degradation of 4-nitrophenol in aqueous medium by electro-Fenton method, J. Hazard. Mater. 145 (2007) 227-232. [17] N. Masomboon, C. Ratanatamskul, M.C. Lu, Chemical oxidation of 2,6-dimethylaniline in the Fenton process, Environ. Sci. Technol. 43 (2009) 8629-8634. [18] J. Anotai, C.C. Su, Y.C. Tsai, M.C. Lu, Effect of hydrogen peroxide on the oxidation of aniline by electro-Fenton and fluid Fenton processes, J. Hazard. Mater. 183 (2010) 888-893. [19] F. Ay, E.C. Catalkaya, F. Kargi, A statistical experiment design approach for the advanced oxidation of azo dye by Direct Red photo-Fenton process, J. Hazard. Mater. 162 (2009) 230-236. [20] N. Masomboon, C.W. Chen, J. Anotai, M.C. Lu, Statistical experimental design for determination of o-toluidine degradation by photo-Fenton process, Chem. Meadow. J. 159 (2010) 116-122. [21] APHA, Standard Methods for the Examination of Water and Wastewater, 18th ed., American Public Health Association, Washington, DC, 1992.

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Journal of Hazardous Materials 196 (2011) 402–411

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Metal Pollution Reconstruction and Recent Sedimentation Processes in Havana Bay (Cuba): A Tool for Coastal Ecosystem Management M. Díaz-Asencio a,∗, J.A. Corcho Alvarado b , C. Alonso-Hernández a , A. Quejido-Cabezas c , A.C. Ruiz-Fernández d, M. Sanchez-Sanchez c, M.B. Gomez-Mancebo c, P. Froidevaux b, J.A. Sanchez-Cabeza and a

Center for Environmental Studies of Cienfuegos, Carretera Castillo de Jagua, Cienfuegos, CITMA-Cienfuegos, Cuba Institute of Radiation Physics (IRA), University Hospital and University of Lausanne, Rue du Grand-Pré 1, 1007 Lausanne Energy Research Center, Swit , Environment and Technology (CIEMAT), Madrid, Spain d National Autonomous University of Mexico. ICMyL, Mazatlán, Mexico and Department of Environmental Science and Technology and Department of Physics, Universitat Autónoma de Barcelona,08193 Bellaterra, Barcelona,

article

i n f o

Article history: Received June 20, 2011 Received in revised form September 12, 2011 Accepted September 12, 2011 Available online September 16, 2011 Keywords: Contaminants 210 Pb 239,240 Pu 137 Cs Dating of sediments Havana Bay

a b s t r a c t Since 1998, the heavily polluted ecosystem of Havana Bay has been the subject of a mitigation program. To determine whether pollution reduction strategies have been effective, we assessed historical pollution trends recorded in bay sediments. The sediment core was radiometrically dated using natural and artificial radionuclides. The anomaly in the 210 Pb record was caused by an episode of accelerated sedimentation. This episode was supposed to happen in 1982, a year that coincides with the heaviest rain recorded in Havana in the 20th century. Peaks in mass accumulation rates (MAR) are associated with hurricanes and heavy rain. Over the past 60 years, these maxima have been associated with ~ periods known to increase rainfall in the northern Caribbean region. We observed a strong steady increase in El Niño pollution (mainly Pb, Zn, Sn and Hg) from the turn of the century to the mid-1990s, with enrichment factors up to 6th MAR and pollution falling rapidly after the mid-1990s, although some levels searchers are still high. This reduction came about thanks to the Integrated Coastal Zone Management program introduced in the late 1990s, which dealt with watershed erosion and pollution. © 2011 Elsevier B.V. All rights reserved.

1. Introduction Havana Bay is one of the largest and most important estuaries on the island of Cuba. However, the economic, economic and recreational values of the bay are threatened by pollution and reduced water depth due to silting [1]. Environmental degradation of the bay ecosystem has intensified in recent decades due to the rapid economic development of the city of Havana. This became the most polluted bay on the island [2]. In order to restore this marine ecosystem, several anti-pollution measures have been implemented over the past decade. However, to assess the impact of environmental management practices, reliable information on pollutant inputs to Havana Bay is needed. In the absence of long-term monitoring data, sediment records can provide retrospective information on past inputs of pollutants to aquatic environments. Contaminants such as heavy metals often have a strong affinity for particulate matter

surfaces and therefore accumulate in sediments. Therefore, dated sediment profiles of major and trace elements can be used to obtain reliable information on the extent and history of pollution and sedimentation conditions [3-7]. Quantitative reconstruction of pollutant input into the water system requires a good sediment chronology. The most widespread method of dating recent sea and lake sediments is based on 210 Pb profiling. The natural radionuclide 210 Pb (T1/2 = 22.23 years) reaches the aquatic environment mainly by precipitation from the atmosphere. However, it can be formed in situ, in the water column and sediments, by the decay of the precursor radionuclide 226 Ra (T1/2 = 1600 years). 210 Pb dating methods are based on the radioactive imbalance between 210 Pb and 226 Ra [8,9]. 210 Pb has proven to be an ideal tracer for dating aquatic sediments deposited in the last 100-150 years, a period of significant environmental change due to industrialization and population growth. 210 Pb dating should always be confirmed with an additional chronostratigraphic marker in the same sediment core [10,11]. Among the most commonly used time markers we find artificially created radionuclides such as 137 Cs, 239, 240 Pu or 241 Am

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ships arriving in the country. Agriculture and intensive forestry in the bay basin have increased soil erosion, and thus the inflow of sediment into the bay. Industrial activities began in the 1850s with the construction of oil refineries, power plants and natural gas production [2,16]. The Havana area experienced rapid economic growth during the 20th century, with a wide range of industries and commercial activities, and a large population growth that required massive urbanization (from 250,000 inhabitants in 1899 to 2.2 million inhabitants in 2001). The uncontrolled development of many activities over the past 400 years has caused serious damage to the natural resources and facilities of Havana Bay. The damage is aggravated by the lack of waste treatment facilities [2,17,18]. The Gulf receives pollutants from many sources, including oil refineries, power plants, municipal and industrial wastewater, three shipyards, river and stream discharges, and atmospheric runoff [1]. 3. Sampling and laboratory methods 3.1. Sampling

Fig. 1. Map of Havana Bay, with sampling station.

[13]. The onset of anthropogenic radionuclides, derived from atmospheric nuclear weapons testing (early 1950s) and their peak value in 1963 [12] have been widely used as time markers in several marine and lagoon studies [13–15] . In this paper, we reconstructed the historical trends of pollutants entering the Havana Bay by analyzing their sediment concentration profiles. The dating of the sediment core was based on the 210 Pb dating method. Due to the low levels of 137 Cs found in the sediments, the date was further confirmed with 239 240 Pu and 241 Am radionuclides. Enrichment factors and pollutants were used to describe the history of pollutants in this aquatic ecosystem. A full geochemical analysis was conducted to examine the potential effects of watershed source change, diagenesis, and atmospheric pollution. The possible origin of the main pollutants and the impact of pollution control measures taken to protect the Havana Bay ecosystem are also discussed. Despite the large number of studies on pollution in coastal environments, few have been conducted in the Caribbean region. Therefore, this study provides important information on pollution trends in coastal ecosystems in this tropical region.

In February 2008, sediment cores were collected by the UWITEC borehole, avoiding the excavation of the bay (Fig. 1). To optimize analytical time and resources, we selected the cores most likely to show a good temporal record (section 1 of the Supporting Information). We took three sediment cores from station B (23◦ 08.107 N 82◦ 20.043 ) at a depth of 8 m (Fig. 1): one core for the analysis of radionuclides, metals and grain size; one for organic contaminants (not listed here) and one frozen for future analysis. The sediment core was extruded vertically and cut into 1 cm pieces. Each part was dried at 45 ◦ C, sieved through a 1 cm sieve and homogenized. The silt content in the sediments showed a slight downward trend from a depth of 15 cm to the surface. The sediments also showed a significant color change in ca. 15 cm deep. 3.2. Laboratory analyzes Grain size was determined by standard methods of analysis with a sieve and a pipette [19]. The content of organic matter for each fraction was calculated by the method of loss on ignition (LOI: 450 ◦ C, 8 h). The total carbon and nitrogen content was measured with a CHN analyzer (LECO TRUSPEC). Total carbon was measured as CO2 with an infrared detector. N2 was measured using a thermal conductivity detector. Inorganic carbon was quantified using an infrared detector (SSM-500, Shimadzu) after sample acidification with phosphoric acid and heating (200°C). Major and trace elements were measured by wavelength dispersive X-ray fluorescence (WDXRF) spectrometry using a rhodium-tube Panalitics System (AXIOS). Total mercury concentrations were determined using an advanced mercury analyzer (LECO AMA254, detection limit 0.01 ng Hg).

2. Description of the place

3.3. Dating of sediments

Havana Bay (NW Cuba) is located outside the city of Havana, it is a typical closed bay with a watershed of about 68 km2. It is characterized by an average depth of 10 m, an area of 5.2 km2 and an average retention time of 7-9 days [1]. The bay is an estuary with deltaic systems in the fluvial outflow zones of the Luyano and Martin Perez rivers and the Tadeo, Matadero, Agua Dulce and San Nicolas streams (Fig. 1). Population density, trade and port activities in the city of Havana have grown significantly since 1850. The city became an important transshipment point between Europe and America in the 19th century. Today, the port of Havana receives about 50%

The dating of the sediment cores was determined by the method (Section 2.3 of the Supporting Information). Sediment samples were placed in sealed plastic containers and stored for at least three weeks to allow the 226 Ra to reach equilibrium with its daughter nuclides. 226 Ra was then analyzed by high-resolution gamma spectrometry using an in-house ORTEC model GX10022 low-background coaxial Ge detector. 226Ra is determined via the 352 keV emitted by its daughter nuclide 214 Pb in equilibrium. The supported 210 Pbsup is derived from the assumption of equilibrium with 226 Ra. The total activity of 210 Pb was determined by high-resolution spectrometry of the ␣ decay product of 210 Po, which is assumed to be 210 Pb

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Fig. 2. Profiles in core B (a) particle size distribution, (b) carbonate and aluminosilicate distribution, and (c) organic (Corg) and inorganic (Cinorg) carbon, phosphorus (P), and nitrogen (N).

to be in balance. Samples (0.5 g) of dry sediment were seeded with 209 Po as a yield indicator and dissolved by adding a 1:1:0.5 mixture of HNO3 + HCl + HF using a microwave analysis system [20]. 210 Po was electronically deposited on silver wafers [21,22], and the measurement was performed on an integrated Camberra alpha spectrometer with ion-implanted silicon planar detectors (450 mm2 active area and 18 keV nominal resolution). 210 Pb in excess (210 Pbex) to 210 Pb supported 226 Ra (210 Pb sup) was determined by subtracting 210 Pbsup from the total 210 Pb activity measured in each layer. 210 Pb ex was then entered into the models to obtain the deposition rate (section 2.3 of the Supporting Information). Measurements of 137 Cs, 239, 240 Pu and 241 Am were used to confirm the 210 Pb dating model. 137 Cs was measured via its emission at 662 keV by high-resolution gamma spectrometry. Sediment samples are then crushed and burned at 550 ◦ C for 48 hours before radiochemical analyzes of 239, 240 Pu and 241 Am. Composite samples were prepared by mixing the layers. The method combines high-pressure microwave digestion for sample dissolution and highly selective extraction chromatographic resins TEVA and DGA (Triskem International, France) for the separation and purification of Pu and Am [23]. Alpha sources were

produced by electrodeposition on stainless steel disks [24]. High-resolution ␣ spectrometry was performed on a ␣ spectrometer with PIPS detectors (Alpha Analyst, Canberra Electronic).

4. Results 4.1. Characteristics of the sediments The sediments are mostly fine and show little variation in grain size in the collected samples (Fig. 2a). Sediments mainly consisted of clay-sized particles (4 m, 15–42%). In the upper 5 cm, the proportion of silt and particles of very fine to coarse sand slightly increased (Fig. 2a). The sediments are mainly composed of carbonates (11-45%) and aluminosilicates (40-80%). The mineral composition of the sediments did not change significantly from the bottom of the core to a depth of 20 cm (Fig. 2b). However, from a depth of 20 cm to the surface, large variations were observed, with the amount of carbonate being negatively correlated with the amount of aluminosilicate (Figure 2b).

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Fig. 3. Profiles of Si, Ca, Fe, Al and Mn in core B. Uncertainties were below 2% for all elements. The dotted line shows the decline.

The amount of carbonate showed a general trend of decreasing toward the surface, but quickly increased to 4 cm (Fig. 2b). The content of inorganic carbon (Cinorg) in the sediments was almost constant throughout the core. but organic carbon (Corg) showed a surface maximum and then decreased with depth, illustrating two zones of rapid change (at 2–3 cm and 16–17 cm depth; Fig. 2c). The high proportion of Corg in the upper 2-3 cm may be related to a change in the source of organic matter, more complex and less biodegradable, typical of industrial organic waste. However, the increase in organic matter may also be related to the decrease in particulate matter observed in the bay in recent decades. In the 3-16 cm segment, Corg was almost constant at around 4%. Then, below 16 cm depth, Corg disintegrated to about 1.5% (Figure 2c). Similar to the Corg sample, nitrogen (N) and phosphorus (P) profiles show increasing trends towards the surface with almost constant concentrations between 3 and 16 cm depth (Figure 2c). 4.2. Major and trace elements in the sediment core Al, Fe, Si and Mn concentration profiles showed a slight decrease in the upper two layers, but no significant changes below a depth of 3 cm (Figure 3). Ca content showed the opposite trend

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in Al (Fe, Si), with slightly higher surface concentrations (Fig. 3). Indeed, Ca, which is probably in the form of biogenic carbonates, acts as a diluent of trace element concentrations in bulk sediments. A small change in the profiles was observed at 15-20 cm, which probably indicates the input of sediments with a high Ca content (Fig. 3). Concentration profiles of Pb, Zn, S, Hg, Sn and Cr in the sediment core show similar increasing trends towards the surface (Figure 4). The maximum concentrations for Zn (450 mg kg−1), Pb (123 mg kg−1) and S (1.4 mg kg−1) were recorded on the core surface, while for Hg (1.4 mg kg−1) , Sn (18.6 mg kg−1) and Cr (365 mg kg−1) maximum concentrations were determined at depths between 5 and 15 cm. The peak concentrations are comparable to those found in other polluted coastal sediments such as Porto Marghera in Italy [25], Halifax Harbor in Nova Scotia [26] and Barcelona in Spain [27]. Similarities in Pb, Zn and Sn profiles (linear correlation R2 > 0.9 and p < 0.01 for each combination) suggest that these elements probably come from the same source and/or have similar geochemical affinities with sediment particles. Pb, Zn, Hg, Sn, Cr and S profiles showed a plateau between 3 and 15 cm depth, indicating a similar time of deposition for the whole section (eg episodic event).

4.3. Radionuclide profiles and sediment chronology 210 The Pbex profile has non-monotonic features that indicate anomalies in the process of sediment accumulation (Fig. 5a). The surface activities of 210 Pbex were around 230 Bq kg−1 (Figure 5a), relatively high compared to activities determined in other studies from Havana Bay [28] and from other Cuban coastal areas [4,29]. Detailed analysis of the distribution of 210 Pbex with depth suggests that the record can be divided into three distinct segments. In the upper (0–3 cm) and lower (15–35 cm) parts of the sediment core, 210 Pb ex decreased exponentially with depth, indicating normal sedimentation. However, 210 Pbex was almost constant throughout the 3-15 cm portion of the core. Smoothing of 210 Pbex refers to either dilution of the atmospheric flux of 210 Pb by sediment mixing, acceleration of sedimentation, and/or occurrence of precipitation due to, for example, heavy rain (frequent in this tropical region). The trends observed in the profiles of 210 Pbex (Fig. 5a), Corg, P, N (Fig. 2c) and some major and trace elements (Figs. 3 and 4) suggest that the 3 to 15 cm section was probably deposited with the same as a result of an episodic event or recession. This is supported by the change in color of the sediment core of about 15 cm

Fig. 4. Profiles (a) of Pb, Zn, Sn and Cr. and b) S and Hg in core B. Uncertainties were: Pb (

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