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This parameter is the preferred spectral comparison method of the widely used NIST mass spectral library search program, and previous work has shown that a threshold of 0. Environmental Protection Agency However, uncertainty of estimated rate constants is not expected to significantly impact calculated reactivity as compounds with the largest contribution to atmospheric reactivity have measured rate constants. Observed and quantified BVOCs include isoprene, two isoprene oxidation products, methyl vinyl ketone and methacrolein, 11 monoterpene species, and two sesquiterpene species McGlynn and Isaacman-VanWertz, Oxygenated species either have volatilities that are too high for efficient trapping e.

Many of the small, oxygenated compounds that might be expected at this site at moderately high abundance e. The exception may be acetaldehyde, which previous work has been shown to contribute non-negligibly to reactivity Hunter et al. A sample chromatogram is shown in Fig. Consequently, we expect that not all sesquiterpenes are captured by this instrument and caution that all reported mixing ratios of sesquiterpenes represent lower bounds.

Periods with gaps are due to instrument issues, and periods reported as zero are below the limit of detection LOD. Many species approached the LOD in the fall and winter months due to low temperatures and decreased incoming shortwave radiation as compared to the warmest months of the year Fig.

Isoprene and its oxidation products Figs. The average mixing ratio of summed isoprene oxidation products was 0. Interestingly, the ratio of isoprene oxidation products to isoprene is variable over the course of the measurement campaign. In addition to differences in their oxidation rates, these differences may be due in part to anthropogenic emissions, both through the influence of NO x on isoprene oxidation pathways and the direct emission of MVK and MACR from vehicles Biesenthal and Shepson, ; Ling et al.

The average mixing ratio of summed monoterpenes in the growing season is 0. Monoterpenes exhibit a similar high period during the growing season but are present throughout the non-growing season as well. The average of summer sesquiterpene mixing ratios in the growing season was 0. Sesquiterpenes, much like monoterpenes, are detected in both the growing and non-growing season.

Average mixing ratios of all classes for each season are provided in Table 2. Diurnal trends in mixing ratios during the growing season May—October and non-growing season November—April are shown in Fig. All terpene classes exhibited the highest mixing ratios in the growing season May—October , when temperature and incoming shortwave radiation were highest Fig. Isoprene and isoprene oxidation products peak in late afternoon hours, as expected due to the light dependence of isoprene Guenther, ; Zimmerman, ; Lamb et al.

Isoprene and oxidation product mixing ratios were typically below the limits of detection in the non-growing season, with little clear diurnal pattern. In contrast to isoprene, summed monoterpenes exhibit peak values in the evening hours, which is consistent with previously reported findings Panopoulou et al. Evening peak values were higher in the growing season than in the non-growing season.

Additionally, daytime lows lasted for longer periods of time in the growing season than in the non-growing season due to longer daylight hours driving more photolytic reactions with the OH radical and shorter lasting nighttime boundary layers Davison et al.

Hourly monoterpene mixing ratios ranged between 0. Sesquiterpene mixing ratios also exhibited peak diurnal mixing ratios in the evening in the growing season. Summed mixing ratios of sesquiterpenes Fig. However, the limit of detection for sesquiterpenes is estimated as 2.

Measured sesquiterpenes therefore provide some insight into the total mixing ratios of sesquiterpenes. Mean sesquiterpene i. Outside the growing season, the mean sesquiterpene mixing ratio was 0. Comparatively, the non-growing season OH reactivity average was 0. While isoprene dominates reactions with OH when present Fig. Due to the year-round presence of monoterpenes, these compounds become the dominant source of OH reactivity in the non-growing season. While some sesquiterpenes may be below level of detection, sesquiterpenes do not generally have OH reaction rates substantially higher than other more dominant terpenes Lee et al.

Ozone reactivity decreases in the non-growing season due to both the decline in isoprene and the decrease in monoterpenes. Average ozone reactivity with isoprene in the growing season is 0. Average ozone reactivity with monoterpenes in the growing season is 2. However, unlike with OH reaction rates, O 3 reaction rates of sesquiterpenes are frequently orders of magnitude larger than dominant monoterpenes, so it is possible that low-abundance, highly reactive sesquiterpenes may still contribute non-negligibly to ozone reactivity.

Important contributions by low-abundance, highly reactive sesquiterpenes have been previously shown in other environments and cannot be excluded by these measurements Yee et al. As in the case of ozone, nitrate reactivity is dominated by monoterpenes due to the slow reaction rates of isoprene and its oxidation products with NO 3. In the non-growing season, isoprene, like isoprene oxidation products and sesquiterpenes, does not contribute to nitrate reactivity.

Monoterpenes dominate nitrate reactivity year-round and have a mean hourly average of 0. Monoterpenes are detected year-round, but small changes in their compositional breakdown i. Figure 8 A breakdown of detected monoterpene isomers in the growing and non-growing seasons for a—b concentration, c—d OH reactivity, e—f ozone reactivity, and g—h nitrate reactivity. Relative contributions from monoterpene isomers are similar for the highest mixing ratio species between the growing and non-growing seasons Fig.

OH reactivity Fig. The stability of the mixing ratios and OH reactivity across a year of measurements suggests that the observed distribution of isomers is a reasonable average representation of monoterpenes in this ecosystem. Instead, a more general quantitative description of the rate at which monoterpenes react with OH or any other oxidant would allow a measurement or estimate of bulk monoterpenes to more accurately be converted into an estimate of their impact on reactivity.

Correlations between calculated monoterpene oxidant reactivity and detected monoterpene mixing ratio Fig. It should also be noted that Fig. The role of structure in atmospheric reactions is even more apparent and critical when considering the reactivity of monoterpenes with ozone. Though the general breakdown of ozone reactivity is qualitatively similar during both the growing and non-growing seasons, there are significant quantitative differences.

The bulk O 3 reaction rate with monoterpenes i. However, while this average rate is relatively stable across seasons, there are periods in the growing season during which the average reaction rate of monoterpenes is substantially faster, which could have impacts during these periods Fig. Isomer dependence of nitrate reactivity is somewhere between O 3 and OH, with an outsize impact of limonene but with a more even split of reactivity across monoterpenes species.

These trends may be explained by the reaction behavior of nitrate. Like the OH radical, nitrate can react with alkenes by either addition to a double bond or abstract a hydrogen, but it has a stronger tendency to add across a double bond, analogous to O 3 Lee et al. These measurements are difficult, however, without robust measurement techniques that do not require significant maintenance.

Using this method, we have collected and are continuing to measure a range of BVOCs in the canopy of a forest representative of the Southeastern US, with periodic coupling of a mass spectrometer to allow for identification of the species of interest. The relative ease of this method gives it great potential for additional long-term BVOC monitoring sites to be set up in more locations.

From this study we have gained a greater understanding of the seasonality of BVOCs ranging from isoprene, isoprene oxidation products, monoterpenes, and sesquiterpenes. Isoprene is important for OH reactivity, but monoterpenes prevail as the most important BVOC class for ozone and nitrate reactivities.

Monoterpenes are observed to be a diverse class of BVOCs with 11 identified compounds detected at the site year-round. This finding is most evident for ozone reactivity but is also the case for OH and nitrate reactivity. The distribution of monoterpenes is qualitatively stable throughout the year, though some important quantitative differences are observed. The bulk reaction rates of the monoterpene class with major atmospheric oxidants presented here therefore provide an improved means to estimate the reactions and impacts of monoterpenes in cases where isomer-resolved measurements are not available e.

DFM conducted the measurement campaign, completed the data analysis, and led the writing of the manuscript. GIVW supervised the study, designed the measurement campaign, and directed the data analysis and writing of the manuscript.

MTL provided feedback on the manuscript. Some authors are members of the editorial board of Atmospheric Chemistry and Physics. The peer-review process was guided by an independent editor, and the authors have also no other competing interests to declare. Tower maintenance and operation were supported in part by the Pace Endowment.

Deborah F. McGlynn and Laura E. The authors gratefully acknowledge the assistance of Koong Yi, Graham Frazier, and Bradley Sutliff in their support in the upkeep and maintenance of the instrument at Pace Tower. We thank Todd Scanlon for the use of the meteorological data provided in Fig.

This paper was edited by Andreas Hofzumahaus and reviewed by two anonymous referees. There is also a generally held opinion that soil washing based on physical separation processes is only cost effective for sandy and granular soils where the clay and silt content particles less than 0.

Soil washing by chemical dissolution of the contaminants is not constrained by the proportion of clay as this fraction can also be leached by the chemical agent. However, clay-rich soils pose other problems such as difficulties with materials handling and solid-liquid separation [ 96 ]. However, where large volumes of soil are to be treated, this cost can be more than offset by reusing clean material on the site therefore avoiding the cost of transport to an off-site centralized treatment facility, and avoiding the cost of importing clean fill.

With physical soil washing, differences between particle grain size, settling velocity, specific gravity, surface chemical behaviour, and rarely magnetic properties are used to separate those particles which host the majority of the contamination from the bulk which are contaminant-depleted. The equipment used is standard mineral processing equipment, which is more generally used in the mining industry [ 91 ]. Mineral processing techniques as applied to soil remediation have been reviewed in literature [ 97 ].

With chemical soil washing, soil particles are cleaned by selectively transferring the contaminants on the soil into solution. Since heavy metals are sparingly soluble and occur predominantly in a sorbed state, washing the soils with water alone would be expected to remove too low an amount of cations in the leachates, chemical agents have to be added to the washing water [ 98 ]. This is achieved by mixing the soil with aqueous solutions of acids, alkalis, complexants, other solvents, and surfactants.

The resulting cleaned particles are then separated from the resulting aqueous solution. This solution is then treated to remove the contaminants e. The effectiveness of washing is closely related to the ability of the extracting solution to dissolve the metal contaminants in soils.

However, the strong bonds between the soil and metals make the cleaning process difficult [ 99 ]. Therefore, only extractants capable of dissolving large quantities of metals would be suitable for cleaning purposes. The realization that the goal of soil remediation is to remove the metal and preserve the natural soil properties limits the choice of extractants that can be used in the cleaning process [ ].

Owing to the different nature of heavy metals, extracting solutions that can optimally remove them must be carefully sought during soil washing. Several classes of chemicals used for soil washing include surfactants, cosolvents, cyclodextrins, chelating agents, and organic acids [ — ].

All these soil washing extractants have been developed on a case-by-case basis depending on the contaminant type at a particular site. Strong acids attack and degrade the soil crystalline structure at extended contact times. For less damaging washes, organic acids and chelating agents are often suggested as alternatives to straight mineral acid use [ ]. Natural, low-molecular-weight organic acids LMWOAs including oxalic, citric, formic, acetic, malic, succinic, malonic, maleic, lactic, aconitic, and fumaric acids are natural products of root exudates, microbial secretions, and plant and animal residue decomposition in soils [ ].

Thus metal dissolution by organic acids is likely to be more representative of a mobile metal fraction that is available to biota [ ]. The chelating organic acids are able to dislodge the exchangeable, carbonate, and reducible fractions of heavy metals by washing procedures [ 94 ].

Although many chelating compounds including citric acid [ ], tartaric acid [ ], and EDTA [ 94 , , ] for mobilizing heavy metals have been evaluated, there remain uncertainties as to the optimal choice for full-scale application.

The identification and quantification of coexisting solid metal species in the soil before and after treatment are essential to design and assess the efficiency of soil-washing technology [ 4 ]. A recent study [ ] showed that changes in Ni, Cu, Zn, Cd, and Pb speciation and uptake by maize in a sandy loam before and after washing with three chelating organic acids indicated that EDTA and citric acid appeared to offer greater potentials as chelating agents for remediating the permeable soil.

Tartaric acid was, however, recommended in events of moderate contamination. Khodadoust et al. Chen and Hong [ ] reported on the chelating extraction of Pb and Cu from an authentic contaminated soil using derivatives of iminodiacetic acid and L-cyestein.

Wuana et al. The use of chelating organic acids—citric acid, tartaric acid and EDTA in the simultaneous removal of Ni, Cu, Zn, Cd, and Pb from an experimentally contaminated sandy loam was carried out by Wuana et al.

These studies furnished valuable information on the distribution of heavy metals in the soils and their removal using various extracting solutions. Phytoremediation, also called green remediation, botanoremediation, agroremediation, or vegetative remediation, can be defined as an in situ remediation strategy that uses vegetation and associated microbiota, soil amendments, and agronomic techniques to remove, contain, or render environmental contaminants harmless [ , ].

The idea of using metal-accumulating plants to remove heavy metals and other compounds was first introduced in , but the concept has actually been implemented for the past years on wastewater discharges [ , ]. Plants may break down or degrade organic pollutants or remove and stabilize metal contaminants. The methods used to phytoremediate metal contaminants are slightly different from those used to remediate sites polluted with organic contaminants.

As it is a relatively new technology, phytoremediation is still mostly in its testing stages and as such has not been used in many places as a full-scale application. However, it has been tested successfully in many places around the world for many different contaminants. Phytoremediation is energy efficient, aesthetically pleasing method of remediating sites with low-to-moderate levels of contamination, and it can be used in conjunction with other more traditional remedial methods as a finishing step to the remedial process.

The advantages of phytoremediation compared with classical remediation are that i it is more economically viable using the same tools and supplies as agriculture, ii it is less disruptive to the environment and does not involve waiting for new plant communities to recolonize the site, iii disposal sites are not needed, iv it is more likely to be accepted by the public as it is more aesthetically pleasing then traditional methods, v it avoids excavation and transport of polluted media thus reducing the risk of spreading the contamination, and vi it has the potential to treat sites polluted with more than one type of pollutant.

The disadvantages are as follow i it is dependant on the growing conditions required by the plant i. Potentially useful phytoremediation technologies for remediation of heavy metal-contaminated soils include phytoextraction phytoaccumulation , phytostabilization, and phytofiltration [ ]. Phytoextraction is the name given to the process where plant roots uptake metal contaminants from the soil and translocate them to their above soil tissues. A plant used for phytoremediation needs to be heavy-metal tolerant, grow rapidly with a high biomass yield per hectare, have high metal-accumulating ability in the foliar parts, have a profuse root system, and a high bioaccumulation factor [ 21 , ].

Phytoextraction is, no doubt, a publicly appealing green remediation technology [ ]. Two approaches have been proposed for phytoextraction of heavy metals, namely, continuous or natural phytoextraction and chemically enhanced phytoextraction [ , ].

Continuous or Natural Phytoextraction Continuous phytoextraction is based on the use of natural hyperaccumulator plants with exceptional metal-accumulating capacity. Hyperaccumulators are species capable of accumulating metals at levels fold greater than those typically measured in shoots of the common nonaccumulator plants. Hyperaccumulator plant species are used on metalliferous sites due to their tolerance of relatively high levels of pollution.

Approximately plant species from at least 45 plant families have been so far, reported to hyperaccumulate metals [ 22 , ]; some of the families are Brassicaceae, Fabaceae, Euphorbiaceae, Asterraceae , Lamiaceae, and Scrophulariaceae [ , ].

Willow Salix viminalis L. A list of some plant hyperaccumulators are given in Table 6. A number of processes are involved during phytoextraction of metals from soil: i a metal fraction is sorbed at root surface, ii bioavailable metal moves across cellular membrane into root cells, iii a fraction of the metal absorbed into roots is immobilized in the vacuole, iv intracellular mobile metal crosses cellular membranes into root vascular tissue xylem , and v metal is translocated from the root to aerial tissues stems and leaves [ 22 ].

Once inside the plant, most metals are too insoluble to move freely in the vascular system so they usually form carbonate, sulphate, or phosphate precipitate immobilizing them in apoplastic extracellular and symplastic intracellular compartments [ 46 ].

Hyperaccumulators have several beneficial characteristics but may tend to be slow growing and produce low biomass, and years or decades are needed to clean up contaminated sites. To overcome these shortfalls, chemically enhanced phytoextraction has been developed. The approach makes use of high biomass crops that are induced to take up large amounts of metals when their mobility in soil is enhanced by chemical treatment with chelating organic acids [ ].

Chelate-Assisted Induced Phytoextraction For more than 10 years, chelant-enhanced phytoextraction of metals from contaminated soils have received much attention as a cost-effective alternative to conventional techniques of enhanced soil remediation [ , ]. When the chelating agent is applied to the soil, metal-chelant complexes are formed and taken up by the plant, mostly through a passive apoplastic pathway [ ].

Unless the metal ion is transported as a noncationic chelate, apoplastic transport is further limited by the high cation exchange capacity of cell walls [ 46 ]. Chelators have been isolated from plants that are strongly involved in the uptake of heavy metals and their detoxification. Enhanced accumulation of metals by plant species with EDTA treatment is attributed to many factors working either singly or in combination.

These factors include i an increase in the concentration of available metals, ii enhanced metal-EDTA complex movement to roots, iii less binding of metal-EDTA complexes with the negatively charged cell wall constituents, iv damage to physiological barriers in roots either due to greater concentration of metals or EDTA or metal-EDTA complexes, and v increased mobility of metals within the plant body when complexed with EDTA compared to free-metal ions facilitating the translocation of metals from roots to shoots [ , ].

Vassil et al. This represents a fold concentration of lead in shoot over that in solution. Since EDTA has been associated with high toxicity and persistence in the environment, several other alternatives have been proposed. Of all those, EDDS [S,S]-ethylenediamine disuccinate has been introduced as a promising and environmentally friendlier mobilizing agent, especially for Cu and Zn [ , , ].

Once the plants have grown and absorbed the metal pollutants, they are harvested and disposed of safely. This process is repeated several times to reduce contamination to acceptable levels. Interestingly, in the last few years, the possibility of planting metal hyperaccumulator crops over a low-grade ore body or mineralized soil, and then harvesting and incinerating the biomass to produce a commercial bio-ore has been proposed [ ] though this is usually reserved for use with precious metals.

This process called phytomining offers the possibility of exploiting ore bodies that are otherwise uneconomic to mine, and its effect on the environment is minimal when compared with erosion caused by opencast mining [ , ].

Assessing the Efficiency of Phytoextraction Depending on heavy metal concentration in the contaminated soil and the target values sought for in the remediated soil, phytoextraction may involve repeated cropping of the plant until the metal concentration drops to acceptable levels.

The ability of the plant to account for the decrease in soil metal concentrations as a function of metal uptake and biomass production plays an important role in achieving regulatory acceptance. Theoretically, metal removal can be accounted for by determining metal concentration in the plant, multiplied by the reduction in soil metal concentrations [ ]. It should, however, be borne in mind that this approach may be challenged by a number of factors working together during field applications.

Prospects of Phytoextraction One of the key aspects of the acceptance of phytoextraction pertains to its performance, ultimate utilization of byproducts, and its overall economic viability. Commercialization of phytoextraction has been challenged by the expectation that site remediation should be achieved in a time comparable to other clean-up technologies [ ].

Genetic engineering has a great role to play in supplementing the list of plants available for phytoremediation by the use of engineering tools to insert into plants those genes that will enable the plant to metabolize a particular pollutant [ ].

A major goal of plant genetic engineering is to enhance the ability of plants to metabolize many of the compounds that are of environmental concern. Currently, some laboratories are using traditional breeding techniques, others are creating protoplast-fusion hybrids, and still others are looking at the direct insertion of novel genes to enhance the metabolic capabilities of plants [ ].

On the whole, phytoextraction appears a very promising technology for the removal of metal pollutants from the environment and is at present approaching commercialization. Possible Utilization of Biomass after Phytoextraction A serious challenge for the commercialization of phytoextraction has been the disposal of contaminated plant biomass especially in the case of repeated cropping where large tonnages of biomass may be produced.

The biomass has to be stored, disposed of or utilized in an appropriate manner so as not to pose any environmental risk. The major constituents of biomass material are lignin, hemicellulose, cellulose, minerals, and ash. It possesses high moisture and volatile matter, low bulk density, and calorific value [ ]. Biomass is solar energy fixed in plants in form of carbon, hydrogen, and oxygen oxygenated hydrocarbons with a possible general chemical formula CH 1.

Composting and compacting can be employed as volume reduction approaches to biomass reuse [ ]. Ashing of biomass can produce bio-ores especially after the phytomining of precious metals. Heavy metals such as Co, Cu, Fe, Mn, Mo, Ni, and Zn are plant essential metals, and most plants have the ability to accumulate them [ ]. Phytostabilization, also referred to as in-place inactivation, is primarily concerned with the use of certain plants to immobilize soil sediment and sludges [ ].

Contaminant are absorbed and accumulated by roots, adsorbed onto the roots, or precipitated in the rhizosphere. This reduces or even prevents the mobility of the contaminants preventing migration into the groundwater or air and also reduces the bioavailability of the contaminant thus preventing spread through the food chain.

Plants for use in phytostabilization should be able to i decrease the amount of water percolating through the soil matrix, which may result in the formation of a hazardous leachate, ii act as barrier to prevent direct contact with the contaminated soil, and iii prevent soil erosion and the distribution of the toxic metal to other areas [ 46 ].

Phytostabilization can occur through the process of sorption, precipitation, complexation, or metal valence reduction. It can also be used to reestablish a plant community on sites that have been denuded due to the high levels of metal contamination. Once a community of tolerant species has been established, the potential for wind erosion and thus spread of the pollutant is reduced, and leaching of the soil contaminants is also reduced. Phytofiltration is the use of plant roots rhizofiltration or seedlings blastofiltration , is similar in concept to phytoextraction, but is used to absorb or adsorb pollutants, mainly metals, from groundwater and aqueous-waste streams rather than the remediation of polluted soils [ 3 , ].

Rhizosphere is the soil area immediately surrounding the plant root surface, typically up to a few millimetres from the root surface. The contaminants are either adsorbed onto the root surface or are absorbed by the plant roots.

Plants used for rhizofiltration are not planted directly in situ but are acclimated to the pollutant first. Plants are hydroponically grown in clean water rather than soil, until a large root system has developed. Once a large root system is in place, the water supply is substituted for a polluted water supply to acclimatize the plant. After the plants become acclimatized, they are planted in the polluted area where the roots uptake the polluted water and the contaminants along with it.

As the roots become saturated, they are harvested and disposed of safely. Repeated treatments of the site can reduce pollution to suitable levels as was exemplified in Chernobyl where sunflowers were grown in radioactively contaminated pools [ 21 ]. Background knowledge of the sources, chemistry, and potential risks of toxic heavy metals in contaminated soils is necessary for the selection of appropriate remedial options.

Remediation of soil contaminated by heavy metals is necessary in order to reduce the associated risks, make the land resource available for agricultural production, enhance food security, and scale down land tenure problems. Immobilization, soil washing, and phytoremediation are frequently listed among the best available technologies for cleaning up heavy metal contaminated soils but have been mostly demonstrated in developed countries.

These technologies are recommended for field applicability and commercialization in the developing countries also where agriculture, urbanization, and industrialization are leaving a legacy of environmental degradation.

Wuana and Felix E. This is an open access article distributed under the Creative Commons Attribution License , which permits unrestricted use, distribution, and reproduction in any medium, provided the original work is properly cited. Article of the Year Award: Outstanding research contributions of , as selected by our Chief Editors.

Read the winning articles. Wuana 1 and Felix E. Academic Editor: A. Received 19 Jul Accepted 23 Aug Published 24 Oct Abstract Scattered literature is harnessed to critically review the possible sources, chemistry, potential biohazards and best available remedial strategies for a number of heavy metals lead, chromium, arsenic, zinc, cadmium, copper, mercury and nickel commonly found in contaminated soils.

Introduction Soils may become contaminated by the accumulation of heavy metals and metalloids through emissions from the rapidly expanding industrial areas, mine tailings, disposal of high metal wastes, leaded gasoline and paints, land application of fertilizers, animal manures, sewage sludge, pesticides, wastewater irrigation, coal combustion residues, spillage of petrochemicals, and atmospheric deposition [ 1 , 2 ]. Fertilizers Historically, agriculture was the first major human influence on the soil [ 21 ].

Pesticides Several common pesticides used fairly extensively in agriculture and horticulture in the past contained substantial concentrations of metals.

Biosolids and Manures The application of numerous biosolids e. Wastewater The application of municipal and industrial wastewater and related effluents to land dates back years and now is a common practice in many parts of the world [ 35 ].

Metal Mining and Milling Processes and Industrial Wastes Mining and milling of metal ores coupled with industries have bequeathed many countries, the legacy of wide distribution of metal contaminants in soil. Air-Borne Sources Airborne sources of metals include stack or duct emissions of air, gas, or vapor streams, and fugitive emissions such as dust from storage areas or waste piles. Lead Lead is a metal belonging to group IV and period 6 of the periodic table with atomic number 82, atomic mass Arsenic Arsenic is a metalloid in group VA and period 4 of the periodic table that occurs in a wide variety of minerals, mainly as As 2 O 3 , and can be recovered from processing of ores containing mostly Cu, Pb, Zn, Ag and Au.

Zinc Zinc is a transition metal with the following characteristics: period 4, group IIB, atomic number 30, atomic mass Cadmium Cadmium is located at the end of the second row of transition elements with atomic number 48, atomic weight Copper Copper is a transition metal which belongs to period 4 and group IB of the periodic table with atomic number 29, atomic weight Mercury Mercury belongs to same group of the periodic table with Zn and Cd. Nickel Nickel is a transition element with atomic number 28 and atomic weight Soil Concentration Ranges and Regulatory Guidelines for Some Heavy Metals The specific type of metal contamination found in a contaminated soil is directly related to the operation that occurred at the site.

Table 1. Soil concentration ranges and regulatory guidelines for some heavy metals. Table 2. Target and intervention values for some metals for a standard soil [ 60 ]. Category Remediation technologies Isolation i Capping ii subsurface barriers. Physical separation Extraction i Soil washing, pyrometallurgical extraction, in situ soil flushing, and electrokinetic treatment. Table 3. Technologies for remediation of heavy metal-contaminated soils. Table 4. Organic amendments for heavy metal immobilization [ 82 ].

Table 5. Inorganic amendments for heavy metal immobilization [ 82 ]. Table 6. Some metal hyperaccumulating plants [ 21 ]. References S. Khan, Q. Cao, Y. Zheng, Y. Huang, and Y.

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Linux - Ubuntu, RedHat, Centos - 64 bit. OpenGL accelerated graphics hardware. Server Java Runtime Environment is required, either version 7 or 8 would work.