Open access peer-reviewed chapter

Organic Waste as Fertilizer in Semi-Arid Soils and Restoration in Mine Sites

Written By

Martha Barajas-Aceves

Submitted: 15 October 2015 Reviewed: 23 February 2016 Published: 30 June 2016

DOI: 10.5772/62665

From the Edited Volume

Organic Fertilizers - From Basic Concepts to Applied Outcomes

Edited by Marcelo L. Larramendy and Sonia Soloneski

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The use of organic waste such as tannery sludge which has high organic matter, N and P content, as organic fertilizer is suitable for improving soil fertility in semi-arid soils and for remediation of abandoned mine sites. Retention of heavy metals on fractional processes of organic matter cannot be generalized, it depends on the chemical characterization of organic waste and soil. Addition of tannery sludge containing high concentrations of Cr and carbonates to semi-arid soils resulted in an increase in Cr loss in infiltration and runoff after 6 months of incubation followed by simulated rainfall. Under these characteristics, results suggest that tannery sludge represents a potential ha ls amended with organic compost. Chemical characteristics of organic waste such as nitrogen content, humified organic matter, pH, EC, CEC, ESP (interchangeable sodium percent), and SAR (sodium absorption ratio) are important properties to consider in organic matter amendment to semi-arid soils participating on the complexity and leaching of heavy metals and nutrients in the matrix of soil.


  • tannery sludge
  • simulated rainfall
  • mobility of heavy metals
  • mine tailings
  • heavy metal fractionation

1. Introduction

Soils of arid and semiarid regions need the amendment of organic compounds due to being low in organic C content and as consequence unable to improve their physicochemical and biological properties and thus their yield of crops and their natural fertility [1, 2].

Organic amendments, such as farmyard manures, have been used as common organic manure for supplying nutrients to plants. After the second world war, there was a great interest in using other organic wastes as organic fertilizer, farmyard manure has been used as a reference when sewage sludge or other organic materials have to be judged [3]. The sewage sludge and sludge compost amendment were ceased when it was found that soluble heavy metals in the soil and the total heavy metals in the crops were increasing when these wastes were applied [4]. However, sewage sludge and compost are the most widely used organic amendments, mainly by their high organic matter, N and P content, which are important nutrients for plant growth, their contribution to soil structure, and resistance to soils erosion [58]. On the other hand, application of sewage sludge as agricultural fertilizer is associated to numerous environmental and health problems such as those implicated by toxic metals, organic compounds, and possible other health problems related to pathogens [912].

The gradual increase in industrialization and urbanization in the last decades has created an enormous increase in volumes of wastes produced all over the world. Usually, these wastes are discharged to the environment [13], especially in countries which regulatory control is not strict.

The objective of the case study described in this chapter, together with methodology and literature review is to provide different scenarios of the role of organic waste in semi-arid soils poor in organic carbon, the effect of heavy metals on soil biogeochemical processes, their dispersion and mobility in soils, and the availability of heavy metals to plants.

1.1. Semi-arid soils and tannery sludge

Tannery sludge derives from a complex combination process where organic and inorganic materials become chemically bound to the protein of the hides and preserve it from deterioration. A significant number of operations within the tannery industry involve large amounts of water, chemicals, and energy leaving as waste large amounts of polluted water. These industrial effluents contain several types of chemicals such as dyes, levelling agents, acids, alkalis, phenols, carbonates, alcohols, cyanide, and heavy metals, among others [14].By-products generated during leather manufacturing are usually rich in proteic matter and organic substances, thus it is a potential resource that can be used as fertilizer in agriculture production. This leather processing waste is an attractive disposal for soil amendment as it has proven to improve the physical properties of soil and supply organic matter and plant nutrients [1519].

The use of these waste in semi-arid soils as organic fertilizer of plants could be an alternative disposal method and simultaneously it will resolve the environmental risks presented when they are abandoned to open sky.

Under natural environmental conditions, chromium is present in either the trivalent Cr(III) or the hexavalent Cr(VI) [20, 21]. The effect of Cr on health have been widely studied [22]. Cr(VI) is about 300 times more toxic than Cr(III).Health effects of Cr have been reported in lung cancer, and birth defects [23, 24].Cr(III) have relatively low toxicity and are easily precipitated and immobilized; however, Cr(VI) is toxic, water soluble, and highly mobile, and can then be transported into the surrounding surface soil and ground water [25, 26].Tannery sludge contains both trivalent (CrIII) and hexavalent (CrVI) chromium. There is little information about the biogeochemical conditions affecting solubility of heavy metals in arid and semi-arid soils including soil pH, Eh, and dissolved organic carbon contents [27]. The information generated in this study is critically important for assessing the benefits or potential risks of using tannery waste to treat semi-arid soils for re-forestation.

1.2. Mine tailings amended with organic wastes

Mine tailing disposal sites from either inactive or abandoned mines are common in arid and semiarid regions throughout the mine region around the world. These tailings have been stored outside and have contaminated local ecosystems and harmed the nearby populations [28]. Today, areas containing mine tailings can be found in urban and agricultural zones and mine tailing storage after the closure of mining operations is becoming increasingly problematic in the arid and semiarid region because of wind erosion. These areas are a source of air pollution giving off particles [29]. Short-term exposure to mine tailing particles can lead to illness, while long-term exposure may lead to premature death in adults and children [30, 31].


2. Research methods

2.1. Sampling areas

The soil was sampled from three sites: Two around mesquite trees (P. laevigata), the dominant vegetation (Dolores Hidalgo, Guanajuato, Mexico), under the canopy and outside the canopy of mesquites (Figure 1), and the third one from a site cultivated with maize (Zea mays) for 20 years.

Figure 1.

Under the canopy tree of mesquite (U), Outside the canopy of mesquite (O).

Soil was collected from a 0- to 5-cm layer, where the highest organic contents can be found. The first sampling was taken from under the canopy of four isolated mesquite trees, 1–2 m from the stem in four perpendicular directions randomly selected. The second one at a distance of 6–8 m from the stem, outside the canopy in the same perpendicular directions (Figure 1). The soil was bulked; all the stones, visible roots, and fauna were removed, it was sieved to less than 2 mm and stored at 5°C to use latter.

2.2. Tannery sludge

Tannery sludge, produced during leather manufacturing, when processing skin or hide to leather, was sampled from a tannery in Leon (Guanajuato, Mexico). It contained large quantities of hair, fatty fleshings, and soluble proteins, as well as sulphide, lime, chromium-sulphate, salts, dyes, acids, and leather trimmings.

2.3. Incubation experiments

A pre-incubation process of soil samples was necessary to allow the soil microbial activity to stabilize after the sampling and sieve management. Soils were pre-incubated for 1 week prior to starting the experiment at conditions similar to the experiment, i.e. at 20°C in the dark, in a temperature and humidity controlled room. Three replicates were destructively harvested at days 0–90 or 120 and stored at −20°C for N mineralization and soil microbial activities analysis.

2.4. Soil microbial activities and nitrification

Maintaining soil fertility depends on biomass and activity of soil microorganisms vital in the biological cycles of most major plant nutrients [32]. Microorganisms are also involved in forming soil structure [33]. Several microbiological parameters have been suggested to measure soil environmental quality [34]. For instance, soil respiration and enzyme activities such as dehydrogenase activity and nitrification, can inform about the presence of viable microorganisms, and on the intensity, kind and time length of the effects of pollutants on the metabolic activity of soils.

Sub-samples of 40 g of soils were placed in 110-ml glass bottles, which were then put into 1-l jars containing 10 ml H2O and a vessel with 20 ml 1 M NaOH solution. The jars were air-tight sealed with plastic lids and incubated at 25 °C for 7 days. After incubation, vessels with 20 ml 1 M NaOH solution were removed, resealed, and stored for future CO2 analysis. At the mentioned dates, soil was removed for analysis of NO3–N, NO2−N, and NH4+−N, done by shaking for 30 min with 100 ml 0.5 M K2SO4 solution and filtered through Whatman No. 42 paper. Similarly, control fresh samples were extracted. Extractants were stored at −20°C until analysis. Concentration of NO3–N and NO2–N in the extracts was determined by colorimetric method [35] and NH4+–N by Indophenol blue [36].

Dehydrogenase activity in soils has been used as measure for overall microbial activity [37]. The method is based on the estimation of triphenyltetrazolium chloride (TTC) reduction rate to triphenyl formazan (TPF) in soils after incubation at 30°C for 24 h. Soil dehydrogenase activity was measured using a modified form of the method used by Casida [37]. Five grams of fresh soil were incubated at 37°C for 24 h in test tubes containing 1 ml 3% 2,3,5- triphenyltetrazolium chloride (TPF), 67 mg CaCO3, and 2.5 ml distilled water. The accumulation of the end-product triphenyl formazan (TPF)was determined in acetone extracts (50 ml) using a PerkinElmer Lamda 3A Spectrophotometer at 520 nm.

2.5. Measurement of CO2-evolved

Glass bottles of 110 ml containing the amended substrate and unamended control soils were placed in 1-l, wide-mouthed glass jars, with a glass flask of 30 ml containing 20 ml of 1 N NaOH solution to trap the evolved CO2. Jars were tightly closed and incubated at 20°C for up to 270 days at room temperature. The NaOH solution was exchanged every 7 days during the first 2 months, and monthly thereafter. Jars were aired each time for 2 min when they were opened to exchange the NaOH solution, to avoid anaerobic conditions in amended and unamended soils. Every time that the NaOH trapped were collected, a blank with non-soils in the jars were collected, too. The values of CO2 in the blanks were used to correct the CO2 trapped inside the jars. CO2 trapped in 1 M NaOH solution was measured in a 5-ml aliquot by titrimetric methods with a standard 0.1 M HCl solution using the phenolphthalein indicator method [38].

2.6. Chemical analysis

Total organic C in the soil and tannery sludge was measured using dichromate digestion [39], total N was measured using Kjeldahl digestion [40], and total hydrolysable and orthophosphate phosphorus were determined using the stannous chloride method [35]. The particle-size was analyzed using the hydrometer method [41]. To conclude, total Cr in tannery sludge, fleshing waste, and infiltration and runoff solutions was measured using absorption atomic spectrometry with a fitted graphite furnace spectrophotometer (Avanta M System 300, GF 3000 S/N 10288). Tannery sludge and fleshing waste were digested with 4:1 HCl: HNO3 using a digiprep TM digestion system, prior to analysis.

2.7. Fractionation of chromium

Tessier et al. scheme [42] is widely used, although application of sequential extraction is still subject controversy. The main problems of sequential extraction procedures are the non-selective use of extracts and the trace elements redistribution among phases during the extraction [43]. In spite of these restrictions, sequential extraction procedures have proved to be useful in the environmental analytical chemistry field [44].

Sequential extraction was utilized for partitioning Cr in soil and sludge amended soils into six operationally defined fractions described by Tessier et al. [42] and modified by Xiong et al. [45]. Six operationally defined fractions, exchangeable, bound to carbonates, bound to Mn oxides, bound to Fe oxides, bound to organic matter and residues according to procedure described by Tessier. Summarizing, 2 g of soil were placed in a 50-ml polycarbonate centrifuge tube and subjected to the following extraction program: Exchangeable fraction (I): soil extracted with 25 ml of 1 M ammonium acetate was shaken for 2 h, then centrifuged. Carbonate bound fraction (II): Fraction I residue extracted with 25 ml of 1 M sodium acetate, adjusted to pH 5 with acetic acid then shaken for 5 h and centrifuged. Mn-oxide-bound fraction (III): Fraction II residue extracted with 25 ml 0.1 M hydroxylamine hydrochloride adjusted to pH 2 with nitric acid then shaken for 12 h and centrifuged. Fe-oxide bound fraction (IV): Fraction III residue extracted with 25 ml of 0.2 M ammonium oxalate, adjusted to pH 3 with oxalic acid, and shaken for 24 h, then centrifuged. Organic and sulphide-bound fraction (V): Fraction IV residue extracted with 5 ml of 30% H2O2 adjusted to pH 2 with HNO3 then heated in a water bath at 85°C. After cooling, 20 ml of I M ammonium acetate were added, shaken for 2 h, then centrifuged. Residual fraction (VI): Fraction V residue digested with 3:1 HCl: HNO3 in digestion glass tubes. After digestion was completed, 25 ml water were added and then filtered. The levels of Cr in the six fractions (I to VI), plus a fresh sample, were analyzed with atomic absorption spectrometry as described above, at 1, 3, and 6 months of incubation.

Mobility factor percentage was calculated according to the equation:


Hexavalent Cr (CrVI) was quantified employing diphenylcarbazide procedure described by Bartlett [46]. One gram of soil was extracted with 3 ml of 10 mM K2HPO4/KH2PO4, pH 7.2. One milliliterazide reagent was added to 8 ml of extract, mixed and stand for 20 min and read the color at 540 nm. Azide reagent was prepared with 120 ml of 85% phosphoric acid, diluted with 280 ml distilled water, to 0.4 g of S-diphenylcarbizide dissolve in 100 ml of 95% ethanol.

2.8. Rainfall system experiment

For the experiment of simulated rainstorm, the rainfall system type Morin [47] was used (Figure 2).

Figure 2.

Rainfall system type Morin.

2.9. Soil treatments

After adjusted to 50% WHC, sub-samples of 10 kg of soil were placed in 15 kg rough use plastic bags. Two different treatments were applied to the soils: tannery sludge (T) to outside (O) and under the canopy (U) of mesquite tree soils, and tannery sludge plus fleshing waste in outside and under the canopy soils (OTF and UTF, respectively). The amount of tannery sludge was chosen to give less than 300 mg Cr kg−1; the critical level of Cr for acceptable use of waste and bio-products in agriculture, as established by the US Environmental Protection Agency [48].

A cotton plug was adjusted to the plastic bags to allow ventilation. After moistening, the treated soils were incubated at 25°C for 1, 3, and 6 months in order to allow mineralization to occur and before they were submitted to a simulated rainfall system. A fresh control was included in order to compare the experimental results with it. The treated soils were air-dried and passed through a 4-mm sieve. They were sampled and then each one was packed to a depth of 4 cm in a 30 × 50-cm box over a 12-cm layer of gravel and 4-cm layer of coarse sand for moisture tension control. Soils were then wetted by capillarity while maintaining the water in a levelled position. Following saturation, the boxes were set at a 3% slope and then allowed to drain for 30 min (Figure 3). A total of four runs for each treated group were conducted, each one after an incubation period (0, 1, 3, and 6 months). For each run, the soil treatments were placed on assoil box carrousel and subjected to a simulated rainstorm using the rainfall system type Morin [47] (Figure 2). The full description of a simulated rain experiment is in Barajas-Aceves et al. [49].

Figure 3.

pH after a simulated rainfall in semi-arid soils amended with fleshing waste and or tannery sludge. Bars indicate standard deviation. Source: Ref. [49].

Simulated rainfall at intensity of 50 mm h−1 and a drop diameter of 3 mm was applied for 60 min. With these specifications, the process simulated the 6 months of a raining period in Dolores Hidalgo, Guanajuato, Mexico. During each run, water percolation through the soil and sediment samples were collected in separated plastic bottles every 5 min for 1 h. Weight of sediment samples and infiltrated water was registered. Double aluminium potassium sulphate (20 ml at 10%) was added to the sediment samples to precipitate suspended solids and separate the water by decantation. The solids in the bottle were dried for 24 h at 105°C and the soil loss was measured by gravimetry, expressed in grams. Once the rainfall was stopped, the soil boxes were left for 24 h to allow the drainage ceased from them, the soil was randomly sampled, allow to dry slowly until it had a moisture equivalent to 40% WHC. Before the rainfall system started, the same soils with the same treatments were sampled and were adjusted to 40% WHC before the soil microbial activities determinations.

The decanted water was transferred to a plastic bottle and weighed. The volume was expressed in cm3 (considering the water density of 1 g cm−3). The runoff was calculated by the equation:


where V is the volume of runoff (cm3), A is the area exposed to rainfall (1500 cm2), and10 is the factor to convert cm to mm.

The infiltration was calculated with a similar equation and expressed in mm [5052].


The results of all these measurements were the sum of the 12 samples per treatment collected in one run.


3. Case study

3.1. Introduction

3.1.1. Semi-arid soils and tannery sludge

Applications of tannery sludge are restricted due to high Cr content, even when Cr(III) content in the sludge is considered to be unavailable. Nevertheless, the use of tannery sludge for reforesting the north of Guanajuato (a natural reserve in Dolores Hidalgo, Guanajuato) is appealing because, it can avoid soil degradation due to the increasing erosion in the region, and at the same time, it can reduce contamination by tannery sludge dumping to open air in Leon, Guanajuato.

The influence of Cr on nitrogen transformation in soils has been studied and reports show somewhat mixed results. James and Bartlett [53] found no inhibition in treatments containing Cr(III) in sewage sludge or tannery effluent but that nitrification was inhibited by Cr(VI) at a concentration of 10 μg g−1 in soil suspensions. Chang and Broadbent [54] observed that nitrogen immobilization, mineralization, and nitrification were inhibited to a great extent by Cr(III) added to a neutral soil, but Cr(VI) was not measured in the extracting solutions used to characterize soil Cr in this study.

Previous studies in semi-arid soils collected in the same natural ecosystem with mesquite amended with tannery sludge to evaluate the biological functioning of soil, show that C and N mineralization increased. Similarly, there was not inhibition in the biological functioning of soil [2].

The aims of this research project were 1) to evaluate the environmental impact of heavy metals from tannery sludge by determining not only the total content in a matrix, but also their bioavailability and their capacity for mobilization and toxicity by chemical fractionation. 2) To indicate the effect of tannery waste type on soil aggregate stability, infiltration, runoff, sediments, nutrients, and Cr loss from semi-arid soils using a simulated rainfall system. This information will be useful for evaluating the ecological risk associated with the use of tannery sludge in semi-arid soils during reforestation projects.

Reducing exposure to contaminants from mining activities is important, especially in the old mining towns like Vetagrande, where large areas have been affected by the presence of mine tailings.

Mine tailings have chemical and physical properties limiting plant growth, like lack of organic matter or macronutrients. Usually, they are acidic, severely toxic, do not have soil structure, have low water retention and slow rates of water infiltration [31]. Organic supplements, such as compost, farmyard manure or biosolids, may be added to overcome these limitations. Moreover, the use of organic compost mixed with tailings could alter several qualities of soil such as potential mineralization of C and N from added organic waste and the mobility and availability of heavy metals in the soil.

The aims of the use of organic amendments on mine tailings were 1) to determine the mobility factor of heavy metals; 2) to evaluate the effect of different organic wastes amended in mine tailings on N and C mineralization potential; 3) to evaluate the availability of Pb and Zn in Brasica juncea as indicators of heavy metal availability for pollutants and two shrubs grown in mine tailings and mixed with compost

3.2. Methods

3.2.1. Sampling area

The soils were sampled from the natural reserve “El Cortijo”, in which the dominant vegetation is mesquite (Prosopis laeviginata), located in Dolores Hidalgo, Guanajuato, Mexico. The soils were collected from two sites: under the canopy and outside the canopy of mesquite tree. The average altitude of sites is between 1750 and 2000 m above sea level and the average annual rainfall is between 400 and 600 mm (mainly from June to September).

3.2.2. Tannery sludge

Tannery sludge produced during leather manufacturing was sampled from a tannery in Leon (Guanajuato, Mexico). The sludge contained large quantities of hair, soluble proteins, and fatty fleshings from processing the skin to hide, and sulphide, lime, chromium-sulphate, salts, dyes, acid, and leather trimmings from processing the hide to leather. Two tannery sludge samples were collected in two different tannery industries with different Cr concentrations. A third tannery sludge was sampled from a tannery in Leon where the fleshing sludge was separated from the leather at the beginning of the process and then followed by chemical treatments. The fleshing waste contained small quantities of hair, soluble proteins and fatty fleshing from the processing of skin to hide. The three different sludge samples and fleshing waste were used in different studies with the same semi-arid soils. Chemical characterization is shown in Table 1. Tannery sludge was air-dried before used for the experimental aerobic incubation.

Tannery sludge 1 Tannery sludge 2 Tannery sludge 3 Fleshing waste (F)
Total C (g kg−1) 281 257.8 7.65 5.68
Total N (g kg−1) 53.4 18.7 0.77 1.5
pH 8.34 8.09 8.65 7.66
Cr (mg kg-1) 6690 1663 6516 136
Na (mg kg-1) 1174.6 119.95
Ca (mg kg-1) 10.94 188.7

Table 1.

Characteristics of tannery waste from Leon Guanajuato, México.

─ = no determined. Data from: Refs. [2, 49, 55].

Amounts of tannery sludge 1 and 2 added to soils were in accord with the criteria of the amount of N covering the recommended dose of N for maize crop growth in the region. The amount of tannery sludge 1 which amended soils from outside and under the canopy of mesquite trees was 1.5 g of wettannery sludge to 50 g of soil. The total amount added to under and outside the canopy soils was approximately 1308 mg C kg−1, 320 mg total N kg−1, 45 mg total P kg−1,and 414 mg Cr kg−1 soils.

The amount of tannery sludge 2 added was 0.0125 g g−1 soils, an amount which covered three times the requirement for the region recommended dose of N for maize crop (i.e. 260 kg N ha−1). The following treatments with three replications were applied at a rate of 250 g g−1 soil: control (without any amendment),Cr(III), Cr(VI), tannery sludge, Cr(III)+ tannery sludge, Cr(VI) + tannery sludge, Cr(III) (Cr2O3), and Cr(VI) (K2Cr2O7). The criteria for dose selected was near to the value of the maximum upper limits of 300 mg kg−1 for acceptable utilization of waste and by-products in agriculture as established by the US Environmental Protection Agency, Part 503 [56].The jars were sealed with air-tight plastic lids and incubated at 25 °C for 180 days. After 0, 30, 60, 120, and 180 days' incubation the CO2 and inorganic N (NH4+, NO2, and NO3) were analyzed as described above.

The amount of tannery sludge 3 (Table 1) was 0.046 g g−1 soils which was equivalent to 22.08 Ton ha−1 and fleshing waste was 0.092 g g−1 soil and was equivalent to 3.7 Ton ha−1. The amount of tannery sludge was applied to give less than 300 mg Cr kg−1 that is the critical level of Cr for acceptable utilization of waste and bio-product in agriculture [48]. These amounts of tannery sludge and fleshing waste contained approximately 13.0 and 2.4% less than 170 kg N ha−1, which is the maximum dose recommended by the European nitrate directive [57] to reduce water pollution by NO3–N from agricultural sources.

3.3. Results and discussion

Addition of tannery sludge 1 to outside and under the canopy of mesquite soils had no inhibitory effect on N mineralization and increased CO2 production and inorganic N concentrations, but did not increase available P concentrations. These results suggest that tannery sludge could provide valuable nutrients to mesquite tree, the dominant vegetation in Dolores Hidalgo [2].

3.3.1. N and C mineralization

The amount of tannery sludge 2 used in another experiment [58]shows that inhibition of nitrification in outside the canopy soils increased when adding tannery sludge plus Cr6+ from 30 to 180 days. Soils under the canopy, amended with the same treatment, did not show a constant effect on nitrification throughout the incubation time (Table 2).

Results from this study showed that nitrification is sensitive to Cr(VI) added alone in outside the canopy soils from 30 to 120 days' incubation (Table 2) and to Cr(VI)plus tannery sludge from 30 to 180 days in soils outside the canopy (data not shown).Cr(III)added alone or Cr(III) plus tannery sludge added to the two soils had no specific effect on the microbial activities (CO2 production or dehydrogenase activity) or N-mineralization [58].

CO2 production rate
Treatment % inhibition
(30 days)
% inhibition
(120 days)
UVI 21.49 25.12
UTVI 7.57 14.71
OVI 98.48
OTVI 22.11
UVI 15.39 15.06
UTVI 29.93 31.48
OVI 61.26 83.63
OTVI 49.61
UVI 52.82
OVI 69.64 83.81
OTVI 62.40 95.36
OTIII 43.38 27.73

Table 2.

Inhibition of CO2 production rate, dehydrogenase activity, and NO3- concentration in under- and outside-the-canopy soils from Dolores Hidalgo, Mexico, incubated at 25°C for 30 and 120 days.

All the values are significant at p ≤ 0.05, ─ not significant. U = under the canopy, and O = outside the canopy soils, VI = Cr(VI), III = Cr(III), T = tannery sludge 2. Source [55].

There is a conflicting effect on using mineralization of organic C in metal contaminated soils because of the stimulation and inhibition on respiration [59, 60]. Results of under the canopy soils show that Cr(VI)could have effects on complex soil organic matter and render it less available by reducing Cr(VI) to Cr(III) [61, 62]. Thus, under the canopy soil which had more organic C, there was less inhibition of CO2 production rate than outside canopy soils.

In soils outside the canopy with low organic matter, Cr(VI) may have effects on soil organic matter available which increases with the dying of cells. Thus, the sum of CO2 produced by death and surviving microorganisms will reflect a small or not inhibition of CO2 production by Cr(VI) (Table 2). Adding tannery sludge plus Cr(VI)may reduce inhibition of CO2 production rate by complexing the Cr(VI) with organic matter from tannery sludge [62].

3.3.2. Cr fractionation

Results from Cr fractionation in under and outside the canopy soils amended with Cr(VI), Cr(III), Cr(VI) plus tannery sludge 2, Cr(III)plus tannery sludge 2, or tannery sludge 2 alone was that the level of total Cr increased in the more resistant fraction (fraction VI which is the least soluble form) and increased further over time. The opposite trend occurred with non-residual fraction (sum of fraction I, II, III, IV, and V) which tended to decrease with time in the two soils (Table 3).

Time incubation 30 days 30 days
Treatments Residual Non-residual Residual Non-residual
Under the canopy
Cr(III) 5.4f 4.1g 41.4c 0.90 h
Cr(VI) 12.7e 31.2f 27.7g 16.8 f
Tannery sludge 26.0d 53.6c 50.1b 25.7 d
Tannery sludge + Cr(III) 31.2c 49.1d 66.4a 33.0 c
Tannery sludge + Cr(VI) 38.4b 74.4b 42.2c 38.0d b
Outside the canopy
Cr(III) 22.0d 7.0g 39.7c 12.0 g
Cr(VI) 25.0d 44.2e 26.5h 24.1 e
Tannery sludge 23.4d 52.2c 33.0 d 32.4 c
Tannery sludge + Cr(III) 45.1a 39.1e 29.7f 26.1 d
Tannery sludge + Cr(VI) 22.0d 77.6a 27.6 g 65.5a

Table 3.

Fractionation of chromium (mg kg1) in semi-arid soils from Dolores Hidalgo, México, amended with Cr(III), Cr(VI) and/or tannery sludge 2, incubated at 25°C for 30 and 120 days.

Residual = fraction VI; Non-residual = sum of fraction I, II, III, IV, and V. Values followed by the same letter in the same column are not significantly different at p ≤ 0.05, according to the Duncan's test.

The greater percentage of Cr in the residual fraction at 120 days of incubation (Table 3) (40–65%) probably reflects the greater tendency of Cr to become unavailable once it is in the soil [63]. The residual and non-residual Cr suggest that the metal bioavailability does not only depend on its concentration but is also affected by the characteristics of the tannery sludge and soil components (such as Fe, Mn, oxides, or the quality of organic matter) into which it is sorbed [61].This will have an impact in the interaction between Cr and the biota [64].

Tannery sludge 3 and fleshing waste were added to the same semi-arid soils incubated for 6 months and subsequently subject to simulated rainfall. In this study we evaluated the Cr loss that occurs due to runoff and infiltration, as well as Cr fractionation, Cr speciation, soil pH, and soil microbial activities before and after the simulated rainfall event.

3.3.3. pH

The highest pH was in soils amended with fleshing waste (OTF > UTF) (average 8.49, 8.12, respectively), followed by OT and UT at 3 months' incubation after rainfall (Figure 3). Similar results were observed in soil before simulated system [49].

The increment in soil pH also increased the solubility of organic molecules, which is especially important in semi-arid soils(Wolf, 1994).The pH decreases from 3 to 6 months agree with those found in a study conducted by Apple [65] who demonstrated that cycles of drying and rewetting might cause disruption of soil aggregation, resulting in a rapid mineralization by soil micro-organisms due to the release of accessible substrate.

3.3.4. Hexavalent chromium

The concentration of Cr(VI) in soils after application of the simulated rainfall was significantly higher in the soils under the canopy treated with tannery sludge and fleshing waste, or tannery sludge only than in soils outside the canopy with the same treatments (Figure 4), even though there were no significant differences between soils treated with tannery sludge or tannery sludge mixed with fleshing waste (data not shown).Values of Cr(VI) remained constant for 1–3 months (1.18–1.74 for outside soils and 2.30–2.80 mg Cr g−1 soil for under the canopy soils) under alkaline conditions (pH > 8.0).

Figure 4.

Chromium hexavalent after a simulated rainfall in semi-arid soils amended with fleshing waste and or tannery sludge. Bars indicate standard deviation. Source: Ref. [49].

3.3.5. Total Cr loss by runoff and infiltration

Total Cr loss in mg was observed in runoff compared to infiltration. The highest values of total Cr released was observed in runoff for UTF treatment, mainly at 1 and 3 months' incubation. The lowest value of total Cr was in infiltration in OTF treatment followed by OT (Table 4) [66].

The high concentration of total Cr observed in runoff from 1 to 3 months corresponds with the highest increase in pH (8.01–8.46). Cr measured between these pH values must be all Cr(VI) where anionic Cr(VI) formations are favoured [67]. Thus, the high concentration of Cr loss in runoff from 1 to 3 months suggests that it might be Cr(VI) which was available to the soil solution. Cr values at 1 and 3 months of incubation for under the canopy soils treated with tannery sludge alone or mixed with fleshing waste were higher than the upper limits of Cr (VI) established by the Mexican Norm [68].

Time (months) OTF UTF UT OT
0 0.42 (0.161) 0.87 (0.074) 1.29 (0.123) 0.50 (0.074)
1 0.09 (0.042) 1.68 (0.134) 0.97 (0.141) 0.29 (0.233)
3 0.40 (0.124) 1.02 (0.128) 2.53 (0.133) 0.30 (0.132)
6 0.34 (0.127) 1.16 (0.187) 0.79 (0.147) 0.50 (0.169)
Total 1.25 4.73 5.58 1.59
0 0.71 (0.156) 1.24 (0.264) 1.33 (0.846) 1.55 (0.285)
1 4.84 (0.150) 11.50 (0.387) 2.64 (0.548) 3.84 (0.402)
3 1.10 (0.380) 13.42 (0.402) 7.89 (0.978) 6.50 (0.151)
6 1.03 (0.191) 1.10 (0.625) 1.01 (0.711) 1.11 (0.205)
Total 7.68 27.26 12.86 13.00

Table 4.

Total Cr (mg) releases in runoff and infiltrations in semiarid soils.

U: soil sampled under the canopy of mesquite tree; O: soil sampled outside the canopy of mesquite tree; T: tannery sludge; F: fleshing waste. Values in parentheses indicate standard deviation. Source [49]. Chromium fractionation

Dominant Cr fractionation in semi-arid soils was bound to carbonates (Fraction II) at 0, 3, and 6 months' incubation after rainfall (Table 5). Dominant Cr fractionation at 1 month incubation was bound to reducible (Fraction III bound to manganese oxide) followed by Fraction IV (bound to Fe oxides). These results demonstrate that Cr binding was influenced by pH and might also have been influenced by ammonium.

0 Months Cr in fraction (%)
OT 0.91 95.30 0 2.98 0.81 0.00
OTF 0.38 95.80 0 2.95 0.83 0.03
UT 0.98 91.79 0 4.64 2.48 0.11
UTF 0.46 92.31 0 4.64 2.47 0.12
1 Month
OT 1.44 4.33 23.51 27.95 42.76 0.02
OTF 2.47 2.07 65.30 17.64 12.51 0.01
UT 3.76 2.20 72.89 12.08 8.82 0.25
UTF 1.64 2.43 78.43 12.22 5.26 0.02
3 Months
OT 2.35 86.77 2.23 2.25 5.99 0.40
OTF 1.90 89.69 2.12 1.57 4.69 0.02
UT 0.47 81.25 3.69 4.01 10.54 0.05
UTF 4.47 29.58 14.20 12.73 38.81 0.22
6 Months
OT 2.91 84.20 5.23 4.11 3.30 0.25
OTF 2.64 84.38 5.57 4.37 2.88 0.16
UT 3.36 83.45 4.90 4.20 4.09 0.01
UTF 2.61 86.41 3.96 3.24 3.78 0.00

Table 5.

Cr Fractionation (%) in semi-arid soils after rainfall amended with fleshing waste (F) and or tannery sludge 3 (T).

Source: Ref.[49].

Fractions: I = exchangeable, II = bound to carbonates, III = bound to Mn oxides, IV = bound to Fe oxides, V = bound to organic matter and VI = residues.

In addition, it has been shown that Cr associated with carbonate becomes susceptible to changes in pH, which results in its becoming soluble [42, 69].

3.3.6. Runoff, soil loss, and infiltration

Water infiltration into soil is one of the most important processes in the hydrological cycle and is crucial in agriculture. Amount of cumulative soil loss and runoff was significantly higher (P<0.01) in soils from outside the canopy than those from under the canopy tree. The opposite occurred in infiltration, being higher under the canopy than outside the canopy tree [66]. Amendment of tannery sludge alone to outside the canopy soils reduced the cumulative soil loss and runoff followed by the tannery sludge plus fleshing waste. However, addition of tannery sludge reduced also cumulative soil loss and runoff under the canopy, but the mixture of tannery sludge plus fleshing waste was higher than those with tannery sludge alone. There were no significant differences in soil infiltration outside the canopy soils with or without amendments. However, infiltration under the canopy soils amended with tannery sludge or tannery sludge plus fleshing waste were higher than without amendment. Thus, the addition of tannery sludge plus fleshing waste to under the canopy soil increased the cumulative runoff [66].

Results reported by Barajas-Aceves et al. [66] suggest that high sodium and salt concentrations of tannery sludge plus fleshing waste (1174.68 mg Na kg−1 tannery sludge and 119.95 mg Na kg−1 fleshing waste and 188.70 mg Ca kg−1 fleshing waste) and soil organic C (0.46 g kg−1 under the canopy soil) affected soil aggregates of under and outside the canopy soils [70] in different ways, thereby allowing reduction of runoff and loss of solids [66, 71]. While elevated electrolyte concentration may enhance flocculation, sodium has the opposite effect in soils, causing dispersion. The contrary behavior occurred with the application of organic C, thus enhancing water infiltration, delaying runoff, and reducing erosion [72]. Thus, organic matter is known to stabilize soil aggregates. Stronger dispersive conditions (such as higher solidicity, lower salinity, and higher energy or impact of water irrigation) should be needed to disperse a stable aggregate [73].

Results of this study could indicate that the degree of dispersion and flocculation and in the treated soils is not only due to salts and sodium concentrations added in the treatments through tannery sludge and fleshing waste, but also to the clay mineralogy of soil [70, 74]. Thus, differences in CEC values for soils under and outside the canopy (51.3 and 22 meq 100 g−1 soil) might reflect the nature of the predominant clay minerals of the soils [75]. Oster et al. [76] studied the flocculation values of montmorillonite and illite suspensions saturated with mixtures of Na and Ca ions in the exchange phase. They suggest that soils with illitic clays are more sensitive to dispersion and clay movement than soils with montmorillonitic clays.

Taken together, these findings suggest that fleshing waste contains sodium at concentrations (data cited above) that adversely impact the infiltration rate of soil collected under the canopy. Previous studies [74, 77] indicate that sodium may cause dispersion and plug soil pores. In this study, repeated wetting and drying of soil could have caused sodium dispersion from sodium coming in the treatments applied to the soil. This, in turn, could have negatively affected the soil structure, reducing infiltration and surface crusting [66, 71, 77].

This would also explain the higher runoff observed in soil collected from under the canopy amended with tannery sludge plus fleshing waste when compared to soil collected from under the canopy and amended with tannery sludge alone. These results suggest that great care must be taken when using waste as an organic fertilizer in semi-arid soil if the wastes contain salts and sodium, to avoid damaging the soil, especially if the goal is to preserve or improve fertile lands. Agassi et al. [74] postulated that the combination of these ESP values and rainwater low salinity caused crust formation and thus high runoff.

3.3.7. Nitrogen loss

The highest values of NO3–N and NH4+–N concentration losses after simulated rainfall were found in the solution infiltration [66].

Outside the canopy of mesquite soils treated with tannery sludge only or tannery sludge plus fleshing waste showed the highest concentrations of NO3–N (341 and 560 mg l−1 respectively) and a lower concentration of NH4+–N (283 and 158 mg l−1 respectively) in the infiltration solution together with the highest reduction of soil loss and infiltration [71] (71 and 55%, respectively) [66] suggesting that no changes regarding nitrification took place in those soils. However, the opposite occurred with soils under the canopy treated with tannery sludge plus fleshing waste, a higher concentration of NH4+–N (452 and 582 mg l−1, respectively) and a lower concentration of NO3–N (445 and 316 mg l−1 respectively) in the infiltration solution were observed. These results, together with no reduction of runoff and low reduction on soil loss (−7 and 43%, respectively) [66], suggest that in areas where treated soil drains poorly, the held water impedes O2 diffusion, creating anoxic areas in the soil at a redox potential appropriate for denitrification. This suggestion is supported by results of high pH (more than 8) in treatments of OTF and UTF at 1 and 3 months of incubation after the simulated rainfall event (Figure 3) [49]. Furthermore, the high values of runoff and low infiltration in UTF amendment [66] suggest that clay mineralogy of under-the-canopy soils and Na concentration, plus the tannery waste added disrupted soil aggregates, leaching the NH4+–N accumulated during the dry intervals prior to the rainfall event [78, 79].

Taken all this information together, it seems that the increasing pH values might have caused competition between chromium oxyanions and OH−, thus decreasing Cr(VI) sorption [80, 81].

Results of the tree different tannery sludge study suggest that tannery sludge characteristics are so important that they define the Cr fractionation in the soil [82]. Thus, the retaining metal in the solid phase depends on metal concentration and abundance of solid phase [83]. Solubility of the metal is mainly controlled by pH, concentration, state of mineral compounds, and type of ligands. Furthermore, these findings indicate that the incubation system used may influence microbial processes in the soil and that the presence of organic compounds plays an important role in determining which fraction of Cr dominates, enhancing metal adsorption to soil phases.

The relative importance of any solid phase for retaining a metal depends on the identity, concentration of metal, and abundance of the solid phase [83, 84]. Solubility of metal is mainly controlled by pH, concentration, type of ligands, chelating agent, oxidation state of mineral component, and the redox potential of the system [85], with TOC enhancing metal adsorption to sediment phases.

The use of tannery sludge with high concentration of carbonates and Cr as organic fertilizer in semi-arid soils could be potentially harmful due to chemical forms related to solubility and carbonate forms favouring heavy metal uptake by plants, and leaching.

3.4. Mine tailings amended with organic wastes

There was a reduction in the mobility of Pb when bokashi and compost of vermicompost were added to agricultural and rangeland soils mixed with mine tailings at 0 or 169 days of incubation. However, mobility of Zn was reduced only in both soils mixed with vermicompost and mine tailings. Differences in Pb and Zn mobility between bokashi and compost treatments might have occurred because the effects of organic materials added to soils on soil properties depend on degradation of such materials, which could have affected heavy metals solubility [86, 87]. The high levels of humified organic matter in vermicompost probably influenced the mobility of Pb and Zn [86]. Some published results suggest that availability of Pb and Zn to soil microflora is also influenced by the high humus content in organic matter [62, 88].

Treatments in agricultural and rangeland soils containing mine tailings plus compost showed the greatest inhibition of cumulative C mineralization followed by bokashi [88]. The highest inhibition of N mineralization in agricultural soils was in treatment amended with vermicompost and in rangeland soils in treatments with compost and bokashi plus mine tailings (20.30 and 18.74% inhibition, respectively). The highest inhibition of dehydrogenase activity was observed in both soil amendments with tannery sludge (48–80% respectively) and the lowest in agricultural soil plus bokashi plus mine tailings and rangeland soils plus vermicompost plus tannery sludge (27 and 39% respectively [88]. These results suggest that the quality of the organic material together with the chemical characteristics of the soils could be important factors influencing decomposition of organic materials. Indeed, contents of nitrogen, cellulose, hemicellulose and lignin, the C/N ratio and the lignin/nitrogen ratio [89] have been reported to be some of the most important factors controlling decomposition processes of organic substrates.

Experimental values of N and C mineralization for the two soils in each treatment [88] were used by fitting to four commonly used kinetic models: Zero order, linearized power function, first order, and first order E [9095].

Nitrogen mineralization in all treatments was best fitted to the linearized power function [88].

N t = K t m aE4

The different values of K and m among treatments and soils did not follow a defined trend. K values in agricultural soils treated with mine tailing alone or with bokashi and vermicompost plus mine tailings increased to 32.1, 26.5, and 31.8% respectively. Similarly, the value of rate constant K in rangeland soils plus mine tailing alone increased to 76.7 % compared to the value of K in soil alone. However, in rangeland soils treated with vermicompost plus mine tailing K decreased to 39%. These results suggest that bokashi and vermicompost in agricultural soils provide a similar pool of mineralizable N and the addition of mine tailings modified the decrease of these pools. Behavior of N mineralization in rangeland soils was different as well as the decrease of N mineralization with the three organic wastes [9092].

The best-fit model for C mineralization was first-order E model for both soils in all treatments [88].

C t = C 0 ( 1 exp kt )+ C 1 bE5

Potential C mineralization (C0) showed the highest values for treatments with low k values [88]. Murwira et al [93] reported that Co*k parameter can be better estimated than only one parameter. Both single parameters are interdependent.

The combination of those two parameters showed high values of Co∗k in treatments with compost, bokashi, or vermicompost alone in agricultural soils, followed by rangeland soils(11.5–18.2 and 8.1–7.6 respectively). These values are in the range of those reported in the literature for different organic materials (40.6–1.3) [95]. Differences in the values of Co∗k in all treatments in both soils might suggest that the amount of lignin-humus present in the organic compost, the amount of nitrogen content, and the organic waste quality and the type of soil might be important in the process and rate of decomposition [96, 97]. The lowest values of Co∗k were in the treatments with mine tailings alone (3.96 in both soils plus mine tailings) compared with soils alone (56 and 65% less in agricultural and rangeland soils, respectively) or the amendment of organic waste alone. These results suggest that the chemical composition of organic waste (nitrogen content, lignin, and polysaccharides) together with the chemical characteristics of the soils could be important factors influencing decomposition of organic materials and even in the presence of heavy metals [98, 99].

Brassica juncea accumulates more Pb and Zn in roots than in shoots [100] growth in soil plus mine tailings. Concentration of Na and Mg measured in mine tailings was 5207 and 102,917 μg g−1 soil respectively [100]. These results demonstrated that Brassica juncea had the ability to survive and tolerate several metals simultaneously, in the presence of high levels of Na and Mg (SAR between 2.5 and 3.7% and ESP between 18.3 to 20.7%) [100]. All metals measured from mine tailings were accumulated in the root and there was very low translocation to the shoots.

According to the literature, the typical behavior of an accumulator species such as Brassica is that there is higher accumulation of heavy metals in the leaves [101], which is opposite to the results of these study.

Results suggest that high levels of Na and salt in mine tailings and the physicochemical characteristic of mine tailings might influence translocation of heavy metals from roots to shoots; metals such as Pb and Zn which are the main metals extracted by Brassica juncea. This suggestion is according to reports showing that salt acts antagonistically, thus when plants grow in media with a high Pb concentration and high salt concentration, the amount of Pb accumulated by Brassica juncea decreases [102].

The mine tailings amended with 10% compost to growth two shrubs Acacia retinodes and Nicotiana glauca were able to survive at high concentrations of heavy metals in mine tailings (Table 1) when 10% compost was added. The dry biomass of both shrubs increase from 62 to 79% growth in mine tailings plus compost compared to mine tailing alone. Echinochloa polystachya was not able to grow on mine tailings, even when it was amended with compost, as was shown by the percentage inhibition data for its root and leaf biomasses. Pb and Zn concentrations in the three plants were higher in roots compared to leaves for all treatments (Pb from 514 to 861 in roots and from 14.6 to 90 (μg g−1) in leaves and Zn from 682 to 766 in roots and 541.4 to 254 (μg g−1) in leaves) [70]. The elevated contents of Pb and Zn in roots along with the low translocation factors [70] indicate that the two shrub species used in this study are appropriate for Pb and Zn phytostabilization.


4. Conclusions

The drastic decrease in soil organic matter in semi-arid soils due to deforestation or in some areas used as deposits for mine tailings are attractive sites for soil restoration. These areas need application of organic residues to avoid subsequent soil erosion by losing soil structure and minimize high risk of pollution in adjacent areas. Application of organic wastes may provide nutrients to pioneering vegetation increasing organic content and improving soil physicochemical and biological properties and thus their natural fertility.

This work provides a clear demonstration of the role of organic waste to increase or release heavy metals according to the quality of organic matter amendment. Recycling valuable components such as C and N available in semi-arid soils was supported by the potential C and N mineralization, dehydrogenase activity, and plant growth. Retention of heavy metals on the fractionation of organic matter cannot be generalized, it will depend on the chemical characterizations of organic waste and soil. Chemical characteristics of the organic waste such as the nitrogen and humified organic matter content, pH, EC, CEC, ESP (exchangeable sodium percent), and SAR (sodium absorption ratio) are important properties to consider in organic matter amendment to semi-arid soils participating on the complexity and leaching of heavy metals and nutrients in the matrix of soil. Measurements such as heavy metal fractionation, percent of heavy metal mobility, soil biochemical processes, and heavy metals accumulation in roots or translocation to plants give a global picture of the complexity of semi-arid soils amendment with organic waste, either in mine soils with heavy metal content, or when used for organic fertilization to semi-arid soils. Behavior of organic composts or organic waste as organic fertilizer in semi-arid soils will find two directions, one: as the organic matter increases heavy metal immobilization in soil and provides nutrients to plants; two: heavy metals present in organic waste or in semi-arid soils from mine sites will interact with soil microbial activities, plant growth, and heavy metal fractionation. Measurement of infiltration factor in both ecosystems containing high heavy metals and salts concentration under simulated rainfall deserve close attention. Monitoring constantly these types of ecosystems to look for the risk of heavy metal mobility after the season changes, and the potential C and N mineralization and heavy metal stabilization in plants should be a priority in semi-arid soils in process of remediation or reforestation.



This work was financially supported by Semarnat-Conacyt, 2002-COI-0184, and FOMIX Zacatecas: ZAC-COI-82037, Mexico. The author would like to thank J. Corona-Hernández for his help in computational tools


  1. 1. Pascual J A, Garcia C, Hernández T, Ayuso M. Changes in the microbial activity of an arid soil amended with urban organic wastes. Biol. Fertil. Soils 1997; 24: 429–434.
  2. 2. Barajas-Aceves M, Dendooven L. Nitrogen, carbon and phosphorus mineralization in soils from semi-arid highlands of central México amended with tannery sludge, Bioresour. Technol. 2001; 77: 121–130. doi:10.1016/S0960-8524(00)00157-7
  3. 3. Johnston A. E. and Wedderburn R. W. M. (1975) The Woburn Market Garden Experiment. 1942-1969. I. A history of the experiment, details of the treatments and the yields of the crops. Rothamsted Experimental Station Report for 1974, Part 2. pp. 79–101.
  4. 4. Le Riche H. Metal contamination of soil in the Woburn Market Garden Experiment resulting from the application of sewage sludge. J. Agric. Sci. 1968; 71: 205–208.
  5. 5. Herrick J E, Wander M. Relationships between soil organic carbon and soil quality in cropped and rangeland soils: the importance of distribution, composition, and soil biological activity. In: Lal R, Kimble J M, Follett R F, Stewart B A, editors. Soil Processes and the Carbon Cycle. CRC Press: Boca Raton, 1997. p. 405–425.
  6. 6. Smoliak S. Effects of manure, straw and inorganic fertilizers on the Northern Great Plains ranges. J. Range Manage. 1965; 18: 11–15.
  7. 7. Marinari S, Masciandaro G, Ceccanti B, Grego S. Influence of organic and mineral fertilizers on soil biological and physical properties. Bioresour. Technol. 2000; 72: 9–17. doi:10.1016/S0960-8524(99)00094-2
  8. 8. Pagliai M, Guidi G, La Marca M, Gichetti M, Lucamante G. Effects of sewage sludges and composts on soil porosity an aggregation. J. Environ. Quality 1981; 10: 556–561.
  9. 9. Barajas-Aceves M, Vera-Aguilar E, Bernal M P. Carbon and nitrogen mineralization in soil amended with phenanthrene, anthracene and irradiated sewage sludge. Bioresour. Technol. 2002; 8: 217–223. doi:10.1016/S0960-8524(02)00151-7
  10. 10. Mc Bride M, Sauve S, Hendershot W. Solubility control of Cu, Zn, Cd, and Pb in contaminated soil. Eur. J. Soil Sci. 1997; 48: 337–346.
  11. 11. Moreno J L, Hernández T, Garcia C. Effects of a cadmium-contaminated sewage sludge compost on dynamics of organic matter and microbial activity in an arid soil. Biol. Fertil. Soils 1999; 28: 230–237.
  12. 12. Wong J W C, Li K, Fang M, Su D C. Toxicity evaluation of sewage sludge in Hong Kong. Environ Int. 2001; 27: 373–380. doi:10.1016/S0160-4120(01)00088-5
  13. 13. United States Environmental Protection Agency (USEPA). Guide for Industrial waste management. EPA 530-R-03-001. USEPA Washington D.C. 2003.
  14. 14. Cooper P. Colour in Dyehouse Effluent. Society of Dyers and Colourists. The Alden Press: Oxford, 1995.
  15. 15. Tester C F. Organic amendment effects on physical and chemical properties of a sandy soil. Proc. Soil Sci. Am. Soc. 1990; 54: 827–831.
  16. 16. Hungalle N, Lal R, Terkuile C H. Amelioration of physical properties by Mucuna after mechanized land clearing of a tropical rainforest. Soil Sci. 1986; 141: 219–224.
  17. 17. Goyal S, Chander K, Mundra M C, Kapoor K. Influence of inorganic fertilizers and organic amendments on soil organic matter and soil microbial properties under tropical conditions. Biol. Fertil. Soils 1999; 29: 196–200.
  18. 18. Aggelides S M, Londra P A. Effects of compost produced from town wastes and sewage sludge on the physical properties of a loamy and a clay soil. Bioresour. Technol. 2000; 71: 253–259. doi:10.1016/S0960-8524(99)00074-7
  19. 19. Bulluck III L R, Brosius M, Evanylo G K, Ristaino J B. Organic and synthetic fertility amendments influence soil microbial, physical and chemical properties on organic and conventional farms. Appl. Soil Ecol. 2002; 19: 147–160. doi:10.1016/S0929-1393(01)00187-1
  20. 20. Narin I, Surme Y, Soylak M, Dogan M. Speciation of Cr(III) and Cr(VI) in environmental samples by solid phase extraction on Ambersorb 563 resin. J. Hazard. Mater. 2006; B136: 579–584. doi:10.1016/j.jhazmat.2005.12.034
  21. 21. Banks M K, Schwab A P, Henderson C. Leaching and reduction of chromium in soil as affected by soil organic content and plants. Chemosphere 2006; 62: 255–264. doi:10.1016/j.chemosphere.2005.05.020
  22. 22. Yassi A, Nieboer E. Carcinogenicity of chromium compounds. In: Nriagu J O, Nieboer E. editors. Chromium in the Natural and Human Environments. John Wiley and Sons: New York, 1988. P. 443–486.
  23. 23. Blasiak J, Kowalik J. A comparison of the in vitro genotoxicity of tri- and hexavalent chromium. Mutation Res. 2000; 469: 135–145.
  24. 24. Morris G S, Guidry K, Hegsted M, Hasten D L. Effect of dietary chromium supplementation on cardiac mass, metabolic enzymes, and contractile proteins, Nutr. Res. 1995; 15: 1045–1052.
  25. 25. James B R, Bartlett R J. Behavior of chromium in soils, VI. Interactions between oxidation-reduction and organic complexation. J. Environ. Qual. 1983; 12: 173–176.
  26. 26. Feldorf S E. Surface reactions of chromium in soils and waters. Geoderma 1995 67: 55–71.
  27. 27. Han F X, Banin A. Solid-phase manganese fractionation changes in saturated arid-zone soils: pathways and kinetics. Soil Sci. Soc. Am. J. 1996; 60: 1071–1080.
  28. 28. Manzanares-Acuña, E, Vega-Carrillo H R, Escobar-León M, Letechipia-de León C, Guzmán-Enríquez L, Hernández-Dávila V, Salas-Luevano M A. Environmental risk assessment of lead in the town of Vetagrande, Zacatecas. Academic Unit of Nuclear Studies, Autonomous University of Zacatecas. 2005. Report UaEN/RI-02-PbVG/INE-02/03081204 (Language Spanish).
  29. 29. Benin A, Sargent J, Dalton M, Roda S. High concentrations of heavy metals in neighborhoods near ore smelters in Northern Mexico. Environ. Health Perspect. 1999; 107: 279–284.
  30. 30. Adriano D.C. Trace Elements in Terrestrial Environments: Biogeochemistry, Bioavailability and Risks of Metals. 2nd Ed. Springer-Verlag, New York, 2001.
  31. 31. Wong M H. Ecological restoration in mine degraded soils, with emphasis on metal contaminated soils. Chemosphere 2003; 50: 775–780. doi:10.1016/S0045-6535(02)00232-1
  32. 32. Robert M, Chenu C. Interactions between soil minerals and microorganisms. In: Stotzky G, Bollang J M, editors. Soil Biochemistry. Marcel Dekker, New York, 1991. p. 307–379.
  33. 33. Harris J A, Birch P. Soil microbial activity in opencast coal mine restoration. Soil Use Manage. 1989; 5: 155–160.
  34. 34. Brookes P C. The use of microbial parameters in monitoring soil pollution by heavy metals. Biol. Fertil. Soils 1995; 19: 341–346.
  35. 35. APHA-AWWA-WPCF. Standard Methods for the Examination of Water and Wastewater. 18th ed. APHA-AWWA-WPCF, Washington, DC, 1992.
  36. 36. Weatherburm M W. Phenol-hypochlorite reaction for determination of ammonia. Anal. Chem. 1967; 39: 971–974.
  37. 37. Casida L E Jr. Microbial metabolic activity as measured by dehydrogenase determination. Appl. Environ. Microbiol. 1977; 34: 630–636.
  38. 38. Zibilske L M. Carbon mineralization. In: Weaver R W, Angle S, Bottomley P, editors. Methods of Soil Analysis. Part 2. Microbiological and Biochemical Properties, Book Series, No. 5. Soil Sci. Soc. Am., Madison, WI, 1994. p. 835–863.
  39. 39. Kalembasa S J, Jenkinson D S. A comparative study of titrimetric and gravimetric methods for the determination of organic carbon in soils. J. Sci. Food Agric. 1973; 24: 1085–1090.
  40. 40. Bremner J M. Total nitrogen. In: Black C A, editor. Methods of Soil Analysis, Vol. 2. American Society of Agronomy, Madison, WI, 1965. p. 1147–1178.
  41. 41. Bouyoucous G J. Hydrometer method improved for making particle size analysis of soil. Agron. J. 1952; 54: 465–466.
  42. 42. Tessier A, Campbell P G C, Bisson M. Sequential extraction procedure for the speciation of particulate trace metal. Anal. Chem. 1979; 51: 844–851.
  43. 43. Bermond A P. Thermodynamics applied to the study of the limits of sequential extraction procedures used for the speciation of trace elements in sediments and soils. Environ. Technol. 1993; 13: 1175–1179.
  44. 44. Nirel P M, Morel F. Technical note: pitfalls of sequential extractions. Water Res. 1990; 24: 1055–1056.
  45. 45. Xiong X, Li P, Stagnitti F, Turoczy N, Allinson G, Sherwood J, LeBlanc M. Speciation of heavy metals in sewage irrigated pastures, Bull. Environ. Contam. Toxicol. 2004; 73: 71–76.
  46. 46. Bartlett R J. Chromium cycling in soils and water: links, gaps, and methods, Environ. Health Perspect. 1991; 92: 17–24.
  47. 47. Morin J, Goldberg, S, Singer I. Rainfall simulator with rotating disk. Trans. Am. Soc. Agric. Eng. 1967; 10: 74–79.
  48. 48. U.S. Environmental Protection Agency. 2002. Using biosolids for reclamation and remediation of disturbed soils. Prepared by the Center for Urban Horticulture, University of Washington. For: Plant Conservation Alliance Bureau of Land Management, US Department of Interior. USEPA (pp. 1–30).
  49. 49. Barajas-Aceves M, Estrada-Santos H, Rios-Berber J D, Oropeza-Mota J L, Rodríguez-Vázquez R. Distribution and mobility of Cr in tannery waste amended semi-arid soils under simulated rainfall. J. Hazard. Mater. 2009; 171: 851–858. doi:10.1016/j.jhazmat.2009.06.087
  50. 50. Mutchler C K, Murphree C E, McGregor K C. Laboratory and field plots for erosion research. In: Lal I, editor. Soil Erosion Research Methods. 2nd Ed. Soil and Water Conservation Society, Ankeny, IA, 1994. p. 11–38.
  51. 51. Mancilla Villa O R, Oropeza Mota J L, Martínez Mena M. Evaluation of bench terraces for commercial forest plantations. Journal of Forest Science. México. 2009; 34: 97–116. ISSN 1405-3586. (Language Spanish).
  52. 52. Betancourt Yáñez P, Oropeza Mota J L, Figueroa Sandoval B, Ordaz Chaparro V, Ortiz Solorio C, Hernández Garay A. Loss of soil and hydrological potential plots with vegetative cover of forage species. Terra. 2000; 18: 263–275. (Language Spanish).
  53. 53. James B R, Bartlett R J. Nitrification in soil suspensions treated with chromium (III VI) salts or tannery wastes. Soil Biol. Biochem. 1984; 16: 293–295.
  54. 54. Chang F H, Broadbent F E. Influence of trace metals on some soil nitrogen transformations. J. Environ. Qual. 1982; 11: 1–4.
  55. 55. Barajas-Aceves M, Corona-Hernández J, Rodríguez-Vázquez R. Chromium fractionation in semi-arid soils amended with chromium and tannery sludge J. Hazard. Mater. 2007; 146: 91–97. doi:10.1016/j.jhazmat.2006.12.001
  56. 56. Harrison E Z, McBride M B, Bouldin D R. The Case for Caution: Recommendation for Land Application of Sewage Sludge and an Appraisal of the U.S. EPA's Part 503 Sludge Rules. Cornell Waste Management Institute, Cornell University, Ithaca, New York, 1999.
  57. 57. European Nitrate Directive. Implementation of Council Directive 91/676/EEC Concerning the Protection of Waters Against Pollution Caused by Nitrates from Agricultural Sources: Synthesis from Year 2000 Member States Reports. Office for Official Publications of the European Communities, Luxemburg, 2002. 44 pp.
  58. 58. Barajas-Aceves M, Ocampo-Velásquez R, Rodríguez -Vázquez R. Effects of Cr3+, Cr6+ and tannery sludge on C and N mineralization and microbial activity in semi-arid soils. J. Hazard. Mater. 2007; 143: 522–531. doi:10.1016/j.jhazmat.2006.09.095
  59. 59. Lighthart B, Baham J, Volk V. Microbial respiration and chemical speciation in metal-amended soils. J. Environ. Qual. 1993; 12: 543–619.
  60. 60. Leita L, De Nobili M, Muhlbachova G, Mondini C, Marchiol L, Zerbi G. Bioavailability and effects of heavy metals on soil microbial biomass survival during laboratory incubation. Biol. Fertil. Soils 1995; 19: 103–108.
  61. 61. Losi M E, Amhein C, Frankenberger W T. Factors affecting chemical and biological reduction of Cr VI in soil. Environ. Toxicol. Chem. 1994; 13: 1727–1735.
  62. 62. Guiller K E, Witter E, McGrath S P. Toxicity of heavy metals to microorganisms and microbial processes in agriculture soils: a review. Soil Biol. Biochem. 1998; 30: 1389–1414.
  63. 63. Lu A, Zhang, S, Shan X. Time effect on the fractionation of heavy metals in soils. Geoderma 2005; 125: 225–234.
  64. 64. Kamaludeen S, Megharaj M, Juhasz A L, Sethunathan N, Naidu R. Chromium–microorganisms interactions in soils: implications to remediation. Rev. Environ. Contam. Toxicol. 2003; 178: 93–164.
  65. 65. Appel T. Non-biomass soil organic N- the substrate for N mineralization flushes following soil drying-rewetting and for organic N rendered CaCl2-extractable upon soil drying. Soil Biol. Biochem. 1988; 30: 1445–1456.
  66. 66. Barajas-Aceves M, Rios-Berber J D, Oropeza-Mota J L, Rodríguez-Vázquez R. Assessment of tannery waste in semiarid soils under a simulated rainfall system. Soil Sediment Contam. An Internat. J. 2014; 23: 954–964. doi:10.1080/15320383.2014.896861
  67. 67. Bartlett R J, James B R. Mobility and bioavailability of chromium in soils. Adv. Environ. Sci. Technol. 1988; 20: 267–304.
  68. 68. Mexican Official Standard NOM-CCA-.31-ECOL, establishes the maximum permissible limits of contaminants in wastewater discharge from industry, agro-industrial, service and treatment of sewage and drainage systems urban or municipal sewer. Official Journal of the Federation. 1993. (Language Spanish).
  69. 69. Ikem A, Egiebor N O, Nyavor K. Trace elements in water, fish and sediments from Tuskegee lake, southeastern USA. Water Air Soil Pollut. 2003; 149: 51–55.
  70. 70. Barajas-Aceves M, Camarillo-Ravelo D, Rodríguez-Vázquez R. Mobility and translocation of heavy metals from mine tailings in three plant species after amendment with compost and biosurfactant. Soil Sediment Contam. An Internat. J. 2015; 24: 223–249. doi:10.1080/15320383.2015.946593
  71. 71. Hardy N, Shainberg I, Gal M, Karen R. The effect of water quality and storm sequential upon infiltration rate and crust formation. J. Soil Sci. 1983; 34: 665–676.
  72. 72. Potter K N, Torbert, H A, Morrison J E. Tillage and residue effects on infiltration runoff and sediment losses on vertisoils. Trans. Am. Soc. Agric. Eng. 1995; 38: 1413–1419.
  73. 73. Shainberg I, Letey J. Response of soils to sodic and saline conditions. Hilgardia. J. Agric. Sci. 1984; 52(2): 1–60.
  74. 74. Agassi M, Shainber I, Morin J. Effect of electrolyte concentration and soil sodicity on infiltration rate and crust formation. Soil Sci. Soc. Am. J. 1981; 48: 848–851.
  75. 75. Ajiboye G A, Ogunwale J A, James T. Characteristics of soils developed over talc at Ode-ogbe, Kogi State, Nigeria. Am.-Eurasian J. Agric. Environ. Sci. 2008; 4: 489–498. ISSN: 1818-6769.
  76. 76. Oster J D, Frenkel H. The chemistry of the reclamation of sodic soils with gypsum and lime. Soil Sci. Soc. Am. J. 1980; 44: 41–45.
  77. 77. Morin J, Benyamini Y, Michaeli A. The dynamics of soil crusting by rainfall impact and the water movement in the soil profile. J. Hydro. 1981; 52: 321–335.
  78. 78. Kremen A, Bear J, Shavit U, Shaviv A. Model demonstrating the potential for coupled nitrification denitrification in soil aggregates. Environ. Sci. Technol. 2005; 39: 4180–4188. doi:10.1021/es048304z
  79. 79. Vourlitis G L, Pasquini S C, Mustard R. Effects of dry season Ninput on the productivity and N storage of Mediterranean-type shrublands. Ecosystems 2009; 12: 473–488.
  80. 80. Rawajfih Z, Nsour N. Thermodynamic analysis of sorption isotherms of chromium (VI) anionic species on reed biomass, J. Chem. Thermodyn. 2008; 40: 846–851.
  81. 81. Wang X S, Li Z Z, Tao S R. Removal of chromium (VI) from aqueous solution using walnut hull. J. Environ. Manage. 2009; 90: 721–729. doi:10.1016/j.jenvman.2008.01.011
  82. 82. [82] Sommers L., Van Vok V, Sopper W E, Giordano P.M. Bastian R. Effects of soil properties on accumulation of trace elements by crops. In: Page A.L., Logan T.S., Ryan J.A. (editors) Land Application of Sludge. Food chain-implications. Lewis Publishers, Chesea, MI. 1987. pp. 5–23.
  83. 83. Ackay H, Oguz A, Karapine C. Study of heavy metals pollution and speciation in Buyak-Menderes and Gediz River Sediments. Water Res. 2003; 37: 813–822.
  84. 84. Gutiérrez M. Trace element concentration patterns in sediments of the Lower Rio Conchos, Mexico. Water Air Soil Pollut. 2000; 121: 259–270.
  85. 85. Enguix González A, Ternero Rodríguez M, Jiménez Sánchez J C, Fernández Espinosa A J, Barragán de la Rosa F J. Assessment of metal in sediments in tributary of Guadalquivir river (Spain). Heavy metal partitioning and Relationship between the water and sediment system. Water Air Soil Pollut. 2000; 121: 11–29.
  86. 86. Shuman L M. Organic waste amendments effect on zinc fractions of two soils. J. Environ. Qual. 1999; 28: 1442–1447.
  87. 87. Almås A R, Singh B R, Salbu B. Mobility of cadmium-109 and zinc-65 in soil influenced by equilibration time, temperature, and organic matter. J. Environ. Qual. 1999; 28: 1742–1750.
  88. 88. Barajas-Aceves M, Rodríguez-Vázquez R. Effects of organic amendments on the mobility of Pb and Zn from mine tailings added to semi-arid soils. J. Environ. Sci. Health, Part B: Pestic. Food Contam. Agric. Wastes. 2013; 48: 226–236. doi:10.1080/03601234.2013.730345
  89. 89. Talbot J M, Treseder K. Interactions among lignin, cellulose, and nitrogen drive litter chemistry—decay relationships. Ecology 2012; 93: 345–354
  90. 90. Seyfried M S, Rao P S C. Kinetics of nitrogen mineralization in Costa Rica soils: model evaluation and pre-treatment effects, Plant Soil 1988; 106: 159–169.
  91. 91. Stanford G, Smith S J. Nitrogen mineralization potentials of soils. Soil Sci. Soc. Am. Proc. 1972; 36: 465–472.
  92. 92. Broadbent F E. Empirical modeling of soil nitrogen mineralization. Soil Sci. 1986; 141: 208–213.
  93. 93. Murwira H K, Kitchmann H, Switft M J. The effect of moisture on the decomposition rate of cattle manure. Plant Soil 1990; 122: 343–353.
  94. 94. Jones C A. Estimation of an active fraction on soil nitrogen. Commun. Soil Sci. Plan. 1984; 15: 23–32
  95. 95. Saviozzi A, Levi-Minzi R, Riffaldi R. Mineralization parameters from organic materials added to soil as a function of their chemical composition. Bioresour. Technol. 1993; 45: 131–135.
  96. 96. Cisternas R E, Yates L R. Nitrogen in litterfall and precipitation and its release during litter decomposition in the Chilean piedmont matorral. Plant Soil 1982; 67: 359–365.
  97. 97. Paul E A, Clark F E. Carbon cycling and soil organic matter. In: Paul E A, Clark F E, editors. Soil Microbiology and Biochemistry. Academic Press: San Diego, CA, 1989. p. 96–98.
  98. 98. Melillo J M, Aber J D, Muratore J F. Nitrogen and lignin control of hard wood leaf litter decomposition dynamics. Ecology 1982; 63: 621–626.
  99. 99. Laishram I D, Yedava P S. Lignin and nitrogen in the decomposition of leaf littering a sub-tropical forest ecosystem at Shiroy hills in north-eastern India. Plant Soil 1988; 106: 59–64.
  100. 100. Barajas-Aceves M, Camarillo-Ravelo D, Juárez-Sánchez F, Rodríguez-Vázquez R. Lead and zinc distribution in Brassica juncea and arid soil amended with mine tailings and bokashi. Fresen. Environ. Bull. 2012; 21: 2026–2037.
  101. 101. Meyers D E R, Auchterlonie G J, Webb R I, Wood B. Uptake and localization of lead in the root system of Brassica juncea. Environ. Pollut. 2008; 153: 323–332. doi:10.1016/j.envpol.2007.08.029
  102. 102. Kadukova J, Papadontonakis N, Naxakis G, Kalogerakis N. Lead accumulation by the salt-tolerant plant Atriplex halimus. International Conference of the Protection and Restoration of the Environment VII, Mykonos, 2004. p. 1–8.

Written By

Martha Barajas-Aceves

Submitted: 15 October 2015 Reviewed: 23 February 2016 Published: 30 June 2016