Of the seventeen elements recognized essential for plant growth, nitrogen (N) has traditionally been considered one of the most important, as it is the most common limiting nutrient (Brady, 1974). Available N exists in most soils in relatively small amounts, while the plant requirements are relatively large (Stevenson, 1982). Annual N supplementation to agricultural soils is commonly practiced in the form of commercial chemical fertilizers or organic residual products such as municipal biosolids, animal manures, or industrial sludges.
Soils are quite often deficient in N, and plants often respond positively to N fertilization (Brady, 1974). Thus, one of the major benefits from biosolids applications can be the addition of N. However, just as with commercial fertilizers, when the N from biosolids applications exceeds the capacity of the plant and soil system to utilize or store it, the available N can transform into nitrate (NO
3
), move through the soil profile, and become a potential groundwater contaminant.
Because biosolids are relatively high in N, the amount of N is usually the limiting factor for applications. A major design criterion for the appropriate application of biosolids is the regulation of the amount of nitrates entering groundwater, especially near aquifers that have the potential for domestic use. The amount of N that can be safely applied to a forest or agricultural crop is determined by the agronomic rate of the crop and the capacity of the soil system to store or utilize the N (EPA, 1994); this amount is both site- and crop-specific. The U.S. EPA has set a drinking water standard of 10 mg L -1 NO
3
-
. This limits the amount of NO
3
which can seasonally move through the soil profile, taking into account dilution within the aquifer.
Nitrogen exists in three basic forms: organic, mineral and gaseous. Most of the N naturally present in the soil and in biosolids are in an organic forms which are not available to plants. Powlson (1993) describes a soil N pool ratio of 25:1 (organic
N:mineral N). The predominant mineral forms of N (ammonium [NH
4
+ ] and NO
3
) are available to plants and are the sole source of N for most plants. Gaseous forms of N are not available to most plants, although some N-fixing plants, such as alfalfa ( Medicogo sp.) and red alder ( Alnus rubra ), can utilize dinitrogen gas (N
2
) from the soil atmosphere.
Nitrogen can be transferred among these forms by biological and chemical processes as outlined in the N cycle in Figure 1.
Most biosolids contain predominantly organic N with a smaller portion of NH
4
+ . The
NH
4
+ is immediately available to plants, but, depending on conditions, it may follow several different pathways other than plant uptake. These include conversion to ammonia gas (NH
3
) and loss to the atmosphere, utilization by soil microorganisms, or conversion to nitrates (Figure 1). The organic N must first be mineralized by soil microorganisms to release the N into a plant-available form. This occurs slowly over a number of years, with

NBMA DRAFT Summary - Nitrogen the majority occurring the first year. The key to the efficient use of biosolids N is the same as for the efficient use of manure or commercial fertilizer N: managing applications to get optimal plant uptake with minimal leaching losses.
2

NBMA DRAFT Summary - Nitrogen
The following sections provide detail on each step of the N cycle, emphasizing the importance of each in managing organic residuals applications. Abstracts of the articles that are referenced in this review are included and are available for further review through the University of Washington.
Initial nitrogen characteristics
Initially, most biosolids contain on average 80 percent organic N, 19 percent NH
4
+ and minimal NO
3
. Sommers (1977) reported a range of biosolids N content from 0.1 percent to 17.6 percent. The forms of N in biosolids will vary according to the material being treated and the treatment method. The dominant form is organic N, which varies from about 50 percent to nearly 100 percent of the total N. Ammonia/ammonium makes up the bulk of the remaining N, while little nitrite (NO
2
) or NO
3
exists, although aerobically digested biosolids will generally have more nitrates present than anaerobically digested biosolids (King, 1984). The work of Wen et al. (1997) indicated that irradiation of three distinct biosolids either had no effect on the NH
4
+ concentration or increased the concentration by 20 or 31 percent. Dewatering generally increases the percentage of N in organic form, while aging or lagooning generally increases the mineral forms (USDA
SCS, 1975). Iwatsubo and Nagayama (1994) found that NH
4
+ accounted for 77 percent of the total N in wastewater applied to a watershed in Japan.
Mineralization
Mineralization (also known as “ammonification”) is the release of NH
4
+ from organic
N by microbial action (Ladd and Jackson, 1982). As discussed above, the majority of N in biosolids exists in an organic form, and mineralization of the N in this form must occur before uptake, other transformations, or losses take place.
Much work has been done on determining mineralization rates for biosolids, particularly during the first year. Common design values may or may not be appropriate.
For instance, typically designers will use a common value of a 20 percent mineralization rate during the first year for anaerobically digested biosolids. However, mineralization studies that have been conducted with municipal biosolids (Epstein et al., 1978; Cassman and Munns, 1980; Magdoff and Amadon, 1980; Terry et al., 1981) point out that a great deal of variability exists. Cripps et al. (1992) reported the results of a field study in which
~50 percent of the organic N provided by biosolids was mineralized in the first year following application; ~30 percent was mineralized in the second year. In a long-term, multiple biosolids application study, Barbarick et al. (1996) determined the first-year mineralization rate following the application of 6.7 dry Mg ha -1 biosolids to range from
13-43 percent, while the rate ranged from 41-67 percent following an application of 26.8 dry Mg ha -1 during the same period. (It should be noted, however, that the lower rate resulted in proportionally less N loss and greater soil accumulation and grain removal of
N.) Mineralization rates of ~50 percent for biosolids and farmyard manure treatments and a rate of 76 percent for a green manure were obtained in a research project discussed by
Hyvönen et al. (1996); their results were generally very similar to those predicted by the
N dynamics model utilized.
3

NBMA DRAFT Summary - Nitrogen
Nitrogen mineralization rates can vary depending upon the characteristics and type of organic residual, the soil type, soil pH, aeration, and climatic conditions such as temperature and moisture (Terry et al., 1981; Serna and Pomares, 1992). Epstein et al.
(1978) found great differences between mineralization rates of raw, digested and composted biosolids. Liquid biosolids showed a higher mineralization rate than composted biosolids according to Chang et al. (1988). Many researchers have found that mineralization rates do not significantly change with application rates (Epstein et al.,
1978; Magdoff and Amadon, 1980; Terry et al., 1981; Simpson et al., 1983), while others have found the opposite effect (Lindemann and Cardenas, 1984; Bockheim et al., in press). Serna and Pomares (1992) performed an incubation study for 16 weeks at 25˚ C and found that aerobically treated biosolids showed a significantly greater N mineralization rate than anaerobically treated biosolids applied at the same rate, with the mineralization rate ranging from 14-46 percent (74-428 mg N kg -1 soil). Sikora and
Yakovchenko (1996) concluded a 2-yr incubation study with the result that the mineralization rate of a biosolids compost+soil was significantly higher than the rate for a co-compost (biosolids+the organic fraction of municipal solid waste [MSW])+soil. In addition, the mineralization rate was significantly lower in compost+sand than in compost+soil. Neither compost increased the mineralization of preexisting soil organic matter as had been expected. The results of an incubation study by Wen et al. (1997) revealed highly significant variability among three anaerobically treated, dewatered biosolids following irradiation. The proportion of N mineralized was more than doubled after irradiation for one biosolids, whereas it was decreased 14 times for another biosolids. Garau et al. (1991) examined the effect on biosolids-borne N mineralization rates due to the addition of a weathered, coal fly ash. Fly ash at two rates and 20 Mg ha -1 biosolids were applied to a soil. It was determined that the addition of fly ash at a rate of
50 Mg ha -1 had an insignificant effect on the mineralization rate, but the rate was significantly decreased when fly ash was applied at a rate of 500 Mg ha -1 . Diez et al.
(1992) recommended that acidic peat be incubated with biosolids to produce a product higher in N (due to an accelerated N mineralization rate) and further suggested liming of the resulting product prior to application. In a laboratory incubation study, Prescott et al.
(1993) showed that forested plots of Douglas-fir ( Pseudotsuga menziesii Mirb.) amended with a total of 142 dry Mg ha -1 biosolids had a significantly lower N mineralization rate in fermentation and humus material than plots amended with commercial N fertilizers and control plots. However, there were no noted significant differences in field incubation studies.
The carbon to nitrogen ratio (C:N) of a given soil amendment may also contribute significantly to the mineralization rate. An incubation study by Lerch et al. (1992) indicated that water-extractable primary amines combined with the C:N ratio were highly correlated with the N mineralization rate. Coutinho et al. (1997) conducted a 90-day incubation study and found a general decrease in N mineralization rates following the addition of increasing rates (0-30 g kg -1 soil) of a composted biosolids (3:1 sawdust:biosolids) with a C:N ratio of ~30:1; however, the decreases were generally at insignificant levels. Sims (1990) found that co-composted biosolids with C:N ratios of
33:1-60:1 applied at 22 Mg ha -1 (250-398 kg N ha -1 ) resulted in negative net N mineralization values (adjusted for inorganic N added with the amendment and
4

NBMA DRAFT Summary - Nitrogen mineralized soil N) ranging from -7.0 to -39.0 mg kg -1 . Iakimenko et al. (1996) reported that treatments with two different biosolids resulted in mineralization rates of 2 times that found in farmyard manure or control samples. This difference might be explained by the
C:N ratio of the manure which was nearly 2 times that of the biosolids, and it was theorized that more of the manure N was already in a mineral form than the N in either biosolids. In a project comparing the mineralization rates due to various biosolids types,
Serna and Pomares (1992) indicated that biosolids with the lowest C:N ratio led to the highest mineralization rates.
Most of the studies discussed above consist of laboratory incubation of biosolids over eight weeks or more with periodic leaching to measure mineral formation of N. These tests are similar to those developed by Stanford and Smith (1972) for mineralizable N in soils. Arguments have been made that studies conducted in the laboratory might not adequately simulate field conditions. Most laboratory studies involve analyses of leachates for mineral N. It is possible that these techniques underestimate mineralization by not accounting for gaseous losses, especially denitrification. The leaching technique potentially creates conditions quite favorable for denitrification: continuous or periods of high moisture content and consistently warm temperatures. From limited field data Fox and Ashley (1982) have suggested that laboratory studies yield mineralization rates that are too low. However, some researchers have suggested just the opposite; lab studies may overestimate mineralization due to ideal incubation conditions.
Nitrogen uptake/growth response
Mineral N in biosolids as it is applied (NH
4
+ or small amounts of NO
3
), or that which is released by mineralization, is available for plant uptake. Most plants can readily absorb both NH
4
+ and NO
3
. Nitrate is the dominant form in well-drained agricultural soils, while NH
4
+ dominates where available N is at a premium and nitrification is low (Olson and Kurtz, 1982). Generally, in agricultural fields which have received frequent N fertilization, the dominant form is NO
3
, while in the temperate forest environments or other unfertilized soils, NH
4
+ is the dominant mineral form.
Numerous N uptake and growth response studies have been conducted on silvicultural species. The vast majority of these studies have found increased uptake and a subsequent growth response enhancement following residual applications. Studies by Chang et al.
(1988) showed total biomass harvested was a function of N input. Liquid biosolids applications to forests 8 years and older increased biomass production in studies done by
Corey et al. (1986). Bockheim et al. (1988) found needle mass of red pines ( Pinus resinosa Ait.) increased in relationship to biosolids application rates. Understory biomass increases of up to 132 percent were measured by Brockway (1983). Moffat and Matthews
(1991) found significant increases in useable timber yields following biosolids applications (as well as a rock phosphate amendment) but found no differences among the various treatments. Additionally, foliar N increased significantly compared to the control and all other treatments in plots receiving a 400 m weight increased significantly following the 400 m
3
3
ha -1 treatment level, and needle
ha -1 level and a biosolids treatment of
200 m 3 ha -1 yr -1 added for 5 consecutive years. Crohn (1995), utilized a biosolids land application model (FORSENTO) and suggested that the length of time required to
5

NBMA DRAFT Summary - Nitrogen produce 244 Mg ha -1 of harvestable wood would be 58, 63, 67, and 80 years for biosolids applied at a rate of 4 Mg ha -1 yr -3 beginning 30 years after harvest, biosolids applied at the same time but at a rate of 8 Mg ha -1 yr -3 , biosolids applied at a rate of 4 Mg ha -1 yr -3 beginning at the initial harvest, and a control, respectively. McDonald et al. (1994) reported that the application of an inorganic fertilizer at a rate of 225 kg N ha -1 resulted in greater height and diameter growth of western red cedar ( Thuja plicata ) after 2 years than the application of 504-610 kg N ha -1 of organic amendments such as biosolids, pulp sludge, fish silage, and mixtures of the four residuals with or without wood ash. After two years, height growth response due to amendment was as follows: inorganic fertilizer > fish silage+ash > biosolids > biosolids+pulp sludge > fish silage+ash+pulp sludge > control > wood ash. With regards to N concentrations in foliage after two years, the inorganic fertilizer, fish silage+ash+pulp sludge and fish silage+ash treatments resulted in significantly higher concentrations than other treatments or the control, and the two biosolids treatments resulted in significantly higher concentrations than ash or the control.
Weetman et al. (1993) performed a similar study of western hemlock ( Tsuga heterophylla ), Pacific silver fir ( Abies amabilis ) and western red cedar plantations treated with biosolids or an inorganic fertilizer. Leader growth of trees in treated plots increased at rates ranging from 0.5 to 3 times those in the control plots during the first growing season. Foliar N concentrations were significantly higher in plots treated with the inorganic fertilizer than those treated with biosolids and significantly higher in biosolidstreated plots compared to controls. Hart and Nguyen (1994) found significant increases in ground level diameter, tree basal area and aboveground biomass following the addition of biosolids-borne N at rates of 500-600 kg ha -1 (~10 Mg ha -1 biosolids) four years after the application. Zabowski and Henry (1994) reported that ponderosa pine ( Pinus ponderosa
Dougl.) foliage N was increased significantly 1 and 2 years following application of ammonium nitrate (NH
4
+ +NO
3
) at 220 kg N ha -1 , but no other significant results were found for this fertilizer during the 5-year study. Over the course of the study, biosolids resulted in the greatest increase among fertilizers tested and was the only fertilizer to result in a significant increase in foliar N during the fifth year following application.
Not all studies of organic amendments have discovered growth rate and/or uptake increases of silvicultural species, however. Polan (1993), found that increasing levels of composted biosolids additions resulted in decreasing growth rates of jack pine due to decreasing levels of available N. Growth rates were compared to those of jack pine seedlings grown in a medium of 80 percent peat and 20 percent vermiculite; the peat was gradually replaced with the composted biosolids in 20 percent increments. Only a 20 percent compost rate resulted in similar growth as the control. Similarly, Aschmann et al.
(1990) reported no increase in hardwood forest overstory species growth nor an increase in N uptake one year following increasing levels of biosolids applications with 0, 200,
400, or 800 kg N ha -1 . During the second growing season, three of five species showed increased N uptake in foliar samples, but only one species showed increased growth due to the increasing biosolids application rates. The authors considered drought to likely have been responsible for the lack of response. However, the data suggested that N uptake by overstory species on plots with the highest biosolids addition rate was ~74 kg N ha -1 greater than on control plots. The addition of two primary pulp and paper sludges led to an increase in soil C:N ratio from 75:1 (control) to 102 and 116:1, respectively; this
6

NBMA DRAFT Summary - Nitrogen resulted in a 32 percent decrease in Douglas-fir height growth after two years (compared to control values) (Harrison et al., 1996). In the same study, Douglas-fir height growth was increased by 146 percent after two years on plots to which secondary pulp and paper sludge with a C:N ratio of 9:1 had been added.
There also have been many examinations of the effects of biosolids and other residuals applications upon N uptake and growth response of agricultural crop species.
Although, there have been conflicting results when biosolids are compared to inorganic fertilizer applications, similar to silvicultural applications, substantial evidence exists that enhanced growth rates and N uptake will follow residual amendments to soils. Menelik et al. (1991) reported that both winter wheat ( Triticum aestivum L.) yield and N uptake were greater following biosolids applications than following inorganic N sources applications at similar N levels. The biosolids additions resulted in 20 percent greater yields and 39 percent greater N uptake. Menelik et al. (1994) found similar results in experiments conducted with corn ( Zea mays ). The work of Cripps et al. (1992) discussed significantly greater corn yields following applications of 11.2 Mg ha -1 of biosolids either incorporated or not and either limed or not compared to a control and plots amended with an inorganic fertilizer. The authors believed that this response on biosolids-amended plots was due to greater N availability and, in years of unusually low rainfall, water retention. Lerch et al.
(1990) discovered significantly greater winter wheat production and quality in biosolidstreated plots compared to plots treated with comparable levels of commercial N fertilizer and control plots during a seven-year project. Summer cabbage ( Brassica oleracea var capitata cv.) was grown in an experiment by Smith and Hadley (1992) to compare the effectiveness of biosolids and other organic forms of N with inorganic forms of N as fertilizers. Cabbage grown on organic residuals-amended plots had a slightly lower growth rate and N foliar concentrations than those grown on plots to which inorganic fertilizer had been applied, but the differences were insignificant. The authors suggested that biosolids-based fertilizers are an effective replacement for inorganic fertilizers and likely result in less negative impacts on the overall soil environment.
Taking into account various factors, research conducted to determine the most appropriate organic residual application rate for achieving a greater growth response/N uptake has revealed mixed results. Fresquez et al. (1990) applied 0, 22.5, 45, or 90 Mg ha -
1 biosolids to partially degraded grasslands. During both growing seasons examined, the three herbaceous species studied each showed significant increases in foliar N; the increase was linear with increasing biosolids applications. An increased growth rate was observed as well, however, the increases were not necessarily linear with biosolids application; applications of 22.5 and 45 Mg ha -1 generally resulted in a better growth response than the higher application rate or the control. Consequently, the authors suggested a biosolids application rate of 45 Mg ha -1 to such ecosystems. Similarly, Devitt et al. (1990) found a significant linear increase in foliar N in tall fescue ( Festuca arundinacea Schreb.) plants with increasing composted biosolids amendment over four months. Growth rate also increased linearly following increasing composted biosolids application rates for a loamy sand and a sandy loam soil but not for a clay soil. The authors suggested that for the latter soil, the CEC was better correlated with growth response than biosolids application rate. Chen et al. (1996) performed a greenhouse study
7

NBMA DRAFT Summary - Nitrogen examining the effect of composted biosolids, composted cattle manure and three inorganic fertilizers upon nutrient uptake and growth of ryegrass ( Lolium perenne L.).
Without inorganic fertilizer additions, N uptake from both the manure and biosolids compost was positively correlated with the application rate (0, 2, 5, or 10 percent of the dry weight of sand); this relationship was more direct with the biosolids compost.
Additionally, cumulative yields of shoots plus roots over four harvests was positively correlated with compost application rate. A general linear increase in winter wheat yield and uptake following increasing biosolids applications ranging from 0-540 kg N ha -1 yr -1 was discussed by Gavi et al. (1997); liming did not affect either variable. Nitrogen utilization on biosolids-applied plots was only 20 percent that on NH
4
+ +NO
3
applied plots. Ritter et al. (1992) found significantly greater growth response from septage applied at a rate of 747 x 10 -5 L ha -1 yr -1 vs. a rate of 66.3 x 10 -4 L ha -1 yr -1 and with lime addition to the septage of 3 g lime L -1 vs. 6 g lime L -1 to corn and fallow fields. Angle et al. (1992) studied the effect of applying varying rates of biosolids-supplied N (0, 78, 156,
312 kg N ha -1 ) upon the growth and N uptake of soybean ( Bradyrhizobium japonicum ).
Foliar N concentration was significantly increased by the 156 and 312 kg N ha -1 rates during the first year. Plant growth was also significantly increased in the first year at the same rates, increasing shoot weights to 18 and 27 g plant -1 , respectively. There were no significant differences during the second year in either uptake or growth response. To examine the effect of humic substances extracts from biosolids and a co-compost (1:1 biosolids:MSW) upon nutrient uptake and growth of barley (species not specified), Ayuso et al. (1996) utilized doses of 0, 2, 5, 10, 50, and 100 mg C L -1 . Nitrogen uptake was greatest with lower doses of humic substances. Likewise, growth of both aerial portions of the barley and roots was decreased with increased dosage; however, no dose levels resulted in significantly lower growth rates compared with the controls. Barbarick et al.
(1996) reported that an application of 6.7 dry Mg ha -1 biosolids matched plant growth needs better than an application rate of 26.8 Mg ha -1 based upon grain production and residual soil N.
A few studies have revealed either no response or decreased growth rate/uptake following organic amendment applications to agricultural lands. For example, an incubation and greenhouse study was conducted by Coutinho et al. (1997) during which biosolids composted with sawdust applied to a sandy soil at rates of 7.5, 15, 22.5, and 30 g kg -1 did not result in increased N uptake or a growth response by ryegrass. However this compost had a C:N ratio of 30:1 and, thus, little available organic N was likely available for microbial activity to promote N mineralization. Additionally, such composts are known to decompose slowly, and the incubation study lasted only 90 days.
A few researchers have investigated uptake of biosolids-supplied N and growth response by horticultural industry plant species. Klock (1997) reported the results of a study where impatiens ( Impatiens wallerana ) and snapdragon ( Antirrhinum majus ) were grown in media of 0, 30, 60, or 100 percent co-compost. The author found no significant differences in final plant concentration of N in either species among any of the media.
However, the best growth response resulted in containers with 100 percent compost.
Bugbee (1996) measured growth of rhododendron ( Rhododendron panticum L. ‘Anah
Kruschke’), arborvitae ( Thujia occidentalis L.), and black-eyed Susan ( Rudbeckia hirta L.
8

NBMA DRAFT Summary - Nitrogen
‘Goldilocks’) due to media pH and the percentage of media as biosolids compost (4:1 yardwaste:biosolids). Media pH had a greater effect on plant growth of rhododendron and arborvitae than percent compost. Black-eyed Susan growth was not affected by pH or compost percentage.
Volatilization of ammonia
Ammonia and NH
4
+ ions are added to the soil with biosolids applications or are produced during biological mineralization of organic N in biosolids. Because NH
3
is a gas at normal atmospheric temperatures and pressures, evolution of NH
3
into the atmosphere is a potential major loss of N under certain conditions. Vogt et al. (1981) suggested that 25-62 percent of N lost after one year from municipal biosolids surfaceapplied to soils could be attributed to gaseous losses, presumably through NH
3 volatilization. Other research has indicated losses of 10 to 60 percent of the original NH
3 from biosolids (Sommers, 1980), 34 to 84 percent from animal wastes (Miner and Smith,
1975) and 10 percent from urea fertilization (Crane, 1972).
There are numerous variables contributing to NH
3
volatilization. Among the variables are: application technique, soil and amendment pH, climatic conditions, and soil cation exchange capacity (CEC). O'Brien and Mitsch (1980) recommended incorporation if the maximum fertilizer value of biosolids is to be realized. King (1973) reported gaseous losses from 16 to 22 percent for incorporated biosolids as opposed to losses from 21 to 36 percent for surface-applied biosolids. Under acidic conditions (pH less than 7) nearly all of the NH
3
is converted to the mineral form, NH
4
+ , and the potential for gaseous losses decreases. Following conversion, NH
4
+ is then sorbed by organic matter or clay particles, effectively taking it out of solution. Above pH 7, more NH
3
is present, increasing the potential for gaseous loss. Koelliker and Miner (1973) identified temperature and wind speed as major factors affecting the transfer coefficient for NH
3
evolving from a surface.
Increased temperature and wind speed resulted in increased NH
3 loss. According to
Freney et al. (1983), a soil’s CEC may be the most important factor affecting NH
3 volatilization.
Immobilization and below-ground N storage
Nitrogen can be stored either temporarily or long-term in the soil by a number of mechanisms: 1) temporary storage on soil exchange sites, 2) immobilization, where micro-biological transformations convert mineral N into organic N, and 3) accumulation in below-ground biomass (plant roots).
The ability of a soil to temporarily store N as NH
4
+ on soil exchange sites is dependent upon the CEC, which will vary with soil properties. This temporary storage allows time for other transformations to occur, such as nitrification, uptake and immobilization. The major contributors to CEC are soil organic material and clay particles. Soil pH also affects CEC, with fewer exchange sites available in more acidic soils.
Studies have shown that N from biosolids applications has been immobilized in the soil (Terry et al., 1981). Gavi et al. (1997) found net N immobilization from biosolids
9

NBMA DRAFT Summary - Nitrogen application rates ranging from 45-180 kg N ha -1
45 kg N ha with NH
4
+
-1
+NO
3
-
one year following application and from
two years after application. These trends were similar following experiments
applications. The authors reported results suggesting that liming may have increased microbial immobilization of N in both the biosolids- and the NH
4
+ +NO
3
amended agricultural plots. Barbarick et al. (1996) found that a biosolids addition rate of
6.7 dry Mg ha -1 resulted in a proportionally greater N immobilization or retention rate of organic N than applications of 26.8 Mg ha -1 .
With regards to immobilization (as with other parts of the N-cycle), the C:N ratio of the amendment seems to be extremely important. Nitrogen can be immobilized in a highly organic material such as an organic forest floor or in a soil through microbial immobilization. This immobilization is a relatively long-term storage of N and generally occurs when the C:N ratio is greater than 30:1. Immobilization can be a major factor in forests but is generally less important in agriculture. However, Sims (1990) studied four co-composts with C:N ratios ranging from 33:1-60:1 and found extensive N immobilization resulting in N deficiency in winter wheat after applications of the cocomposts at rates of 0, 11, 22, and 44 Mg ha -1 (250-398 kg N ha -1 ) even with the addition of 100 mg kg -1 of NH
4
+ +NO
3
. Amendments to biosolids such as sawdust tend to immobilize N, thereby reducing its loss as a gas and delaying the formation of nitrates
(Vogt et al., 1981). This may be the case due to a general increase in the C:N ratio with increasing levels of woody additions. This effect was also reported by King (1984). Wood residue was shown to immobilize excess available N and subsequently decreased rates of
NO
3
leaching with increasing wood bark content (Sabey et al., 1975). Epstein et al.
(1978) reported that wood chips added to biosolids decreased mineralization rates. The authors theorized that the cause of the decreased mineralization was a high C:N ratio that immobilized large quantities of N. The various ratios of wood to biosolids were studied again by Sabey et al. (1977) to determine the effect on wheat growth. In almost all cases, growth increased with increasing ratios of biosolids to wood. This was presumably due to a decreased N immobilization (and subsequently greater available N) in the higher biosolids:sawdust ratio applications.
Below-ground storage in biomass can also be a significant N sink in forests. Fine roots were estimated to require over 73 and 44 kg N ha -1 yr -1 for midwest hardwood and red pine stands, respectively (McClaugherty et al., 1982). Schiess and Cole (1981) measured over 196 kg N ha -1 yr -1 of below ground accumulation of N from wastewater additions to poplar ( Populus sp.) and Douglas-fir plots during the first five years following planting.
Nitrification
Unless specifically used as a fertilizer or as a product of atmospheric deposition, nitrates exist primarily as a result of nitrification (Schmidt, 1982). Nitrification is the biological transformation of NH
4
+ to NO formation of NO
2 of NO
2
to NO
3
-
3
by a two step process. The first step is the
, usually by the bacteria Nitrosomonas, followed by the transformation
, usually by the bacteria Nitrobacter.
10

NBMA DRAFT Summary - Nitrogen
Both steps of the nitrification process are acidifying, i.e., release H + ions; thus one way of controlling nitrification is by controlling soil pH. Wells et al. (1986) reduced nitrification in biosolids-amended soils by soil acidification to pH 6.6. Conversely, Terry et al. (1981) found nitrification was faster in soils receiving biosolids than in controls because of an increase in pH from biosolids addition.
Numerous other factors also appear to affect rates of nitrification. Nitrification was found to be much greater when moisture was not limiting (Terry et al., 1981). Also limiting nitrification was high concentrations of soil atmosphere CO
2
(Keeney et al.,
1985), such as those that could exist when a lot of organic material is added to the soil. A model simulation was developed by Otoma and Kuboi (1985) which was applied for estimating the daily net rate of NO
3
formation. Results indicated that the rate was accelerated by rain and slowed abruptly soon after. Burton et al. (1990) suggested that nitrification was influenced by biosolids type, forest type and the biosolids-application history of a site. The authors derived the following conclusions: the organisms responsible for nitrification were contained within the biosolids; only an anaerobically digested biosolids resulted in net nitrification over the 8-week incubation period; incorporating the biosolids into the soil reduced net nitrification compared to surface applications; and elevated nitrification rates were found from cores that had received biosolids 33 months prior, suggesting that nitrification could be greater in soils to which biosolids had previously been applied. However, the authors also suggested that incubation studies (such as this study) were not consistent predictors of the NO
3
concentrations in soil leachate and groundwater and that field studies were a more appropriate means of examination. Nine years following the cessation of applications,
Iwatsubo and Nagayama (1994) reported a nitrification rate higher than control values in a watershed that had received a total input of 724 kg N ha -1 over 3 years. According to
Aschmann et al. (1992), the net nitrification rate was greater for forested plots amended with 400 kg N ha -1 than on plots amended with 800 kg N ha -1 . Shepherd (1996) added a nitrification inhibitor, Dicyandiamide, with applications of liquid biosolids, and discovered a decrease in nitrification; a resultant decrease in N loss was detected.
Denitrification
Other than plant uptake (discussed earlier), the fates of nitrates are potentially only two: denitrification, and leaching. Firestone (1982) identified denitrification as the major biological process through which soil N is lost to the atmosphere. The general requirements for denitrification are: 1) the presence of denitrifying bacteria, 2) the presence of decomposable organic matter, 3) anaerobic conditions, and 4) the presence of
N-oxides. The two major potential products of denitrification are N
2
O and N
2
.
Often interest in denitrification stems from N balances made from field measurements, such as N application rates, NO
3
leaching, uptake, and changes in soil N, which result in large portions of N unaccounted for. Bockheim et al. (in press) found 50 to 59 percent of the applied N from a combined pulp and paper sludge could not be accounted for by storage in the trees, sludge/forest floor complex and upper 27 cm of mineral soil, or by leaching beyond the rooting zone. They attributed the imbalance to denitrification losses. Since sludge and soil pH were fairly low, denitrification seemed to
11

NBMA DRAFT Summary - Nitrogen be a more likely pathway than NH
3
volatilization. Brockway (1983) was also unable to account for 51 percent of papermill sludge N and biosolids N applied to a forest, which also suggests significant denitrification. King (1973) measured 20 percent of applied N from municipal biosolids lost due to denitrification. Myrold (1988) found that on an annual basis, the denitrification rate on plots with winter wheat was nearly three times the rate on plots of ryegrass ( Lolium multiflorum Lam.) (1.7 kg N ha -1 yr -1 vs. 0.7 kg N ha -1 yr -1 , respectively). The author concluded that the denitrification rate was positively correlated with soil water content and soil respiration rate, somewhat negatively correlated with soil temperature, and not correlated with the soil NO
3
level. Overall nitrification rates in this study, however, were quite low. Conversely, in a study by
Artiola and Pepper (1992a), the denitrification rate following biosolids amendment was highly correlated with the NO
3
level in soil cores over a 10-day incubation study, with the maximum percentage of NO
3
denitrified ranging from 15-22 percent (compared to 1-
4 percent under “natural” conditions). Additionally, this study found that the denitrification rate for biosolids- and inorganic N fertilizer-amended plots was significantly greater within the incorporation zone (0-30 cm) than at lower depths.
Leaching of ammonium
Because the NH
4
+ ion has a positive charge it is not as susceptible to leaching as anions due to soil properties. Relatively little NH
4
+ has been seen leaching from forest soils applied with biosolids (Hornbeck et al., 1979), or from N fertilization (Crane, 1972).
Insam and Merschak (1997), however, found significant differences in NH
4
+ in soil water in mixed species and spruce ( Picea sp.) forest soil columns during weeks 3-10 of a 29week incubation study for a biosolids+shale treatment and in the spruce forest for a biosolids+bark (2:1 ratio) amendment during portions of the study.
General
Nitrate leaching is the movement of the oxidized ionic form of N (NO
3
) downward through the soil below the rooting zone eventually entering into either groundwater or surface water. The amount of leaching depends primarily upon the application rate, the soil type and climatic conditions, among other variables. Artiola (1991) suggested that, in addition to the amount of N applied, water management practices, specifically irrigation efficiency, (as well as soil texture and position in the profile) also are important in determining potential leaching. High amounts of nitrates in groundwater can be harmful to cattle and to infants less than six months old. The U.S. EPA has established a safe drinking water standard of 10 mg L -1 NO
3
. High nitrates are generally not a concern to older children and adults. Surface water may also have excessive nitrates; studies by de
Haan (1986) showed N in drainage water to be almost entirely in the NO
3
form. High nitrates which have leached from a site and entered surface waters also can promote growth of algae and weeds. The key to managing biosolids applications is keeping nitrates at acceptable levels.
12

NBMA DRAFT Summary - Nitrogen
When N added by humans or N-fixing plants is too abundant in the system, NO
3
formation and subsequent leaching is almost a certainty. Leaching of nitrates has been identified by many as a limiting factor for application of both municipal biosolids and industrial sludge applications (Breuer et al., 1979; Sidle and Kardos, 1979; Sommers,
1980; Riekerk, 1982; Brockway and Urie, 1983; Henry and Cole, 1983; Ashton et al.,
1985).
Forest applications of biosolids
As stated above, the application rate is of primary concern when determining the potential for NO
3
leaching from forested ecosystems. Zasoski et al. (1984) and Riekerk
(1978) have summarized many studies conducted at the University of Washington's Pack
Forest. Heavy applications of up to 25 cm of dewatered biosolids (470 dry Mg ha -1 ) to forested areas resulted in concentrations of nitrates in the soil solutions of up to 500 mg
L -1 . Applications of 24, 47, 94 and 188 Mg ha -1 of biosolids (900-7000 kg N ha -1 ) were estimated to result in 0, 68, 358 and 1052 kg ha -1 NO
3
leaching, respectively (Henry and
Cole, 1986). By utilizing the N dynamics model, FORSENTO, Crohn (1995) found that a biosolids application rate of 8 dry Mg ha significant increases of NO
3
-
-1 yr -3 for 50 years would likely lead to
in groundwater, but an application rate of 4 Mg ha -1 yr -3 would likely not have a significant effect. Similarly, Crohn and Haith (1994) utilized
FORSENTO to predict that the application of 4.5 dry Mg ha biosolids (225 kg N ha -1 yr -3
-1 yr -3 aerobically digested
) would not lead to excessive (>10 mg L -1 ) NO
3
leaching in
99 percent of all years required for successful timber growth. To maintain NO
3
levels in percolation water of <10 mg L -1 , Haith et al. (1992) suggested a range of 0.7-1.5 to 4.2-
5.1 Mg ha -1 yr -1 biosolids addition to New York conifer and hardwood forests based upon worst- and best-case scenarios. Applications of 25 Mg ha -1 to northeastern hardwood forests resulted in minor elevations in NO
3
in soil water, from below 0.5 to 2-3 mg L -1
(Hornbeck et al., 1979). Concentrations of nitrates greater than 10 mg L -1 were seen following applications of 27 and 14 dry Mg ha -1 to mixed hardwood forests in
Pennsylvania (Sidle and Kardos, 1979), with maximum soil water concentrations reaching 194 and 290 mg L -1 , respectively. A computer simulation model based on a solution to the convection diffusion equation was applied to their field experiment data.
The data strongly indicated that soluble NO
3
was being channeled through interconnected macropores in the soil profile. In Europe, it is recommended that biosolids application rates to pine plantations not exceed 9.5 dry Mg ha -1 undigested biosolids or 16 dry Mg ha -1 digested biosolids to prevent leaching of NO
3
(Ahtiainen, 1986). Loading rates of 402 kg N ha -1 caused no adverse affects to soil or groundwater in experiments reported by Corey et al. (1986). Wells et al. (1986) determined that application rates of
800 kg N ha -1 of high N biosolids caused leaching of 22 percent of the N as NO
3
meter depth, but levels declined in 18 months to near 10 mg L -1
at the 1
. Hart and Nguyen (1994) reported that applications of biosolids with 500-600 kg N ha -1 to bigtooth aspen ( Populus gradidentata Michx.) stands, resulted in significant, but environmentally acceptable levels of NO
3
in groundwater. Soil solution NO
3
averaged 11.53, 2.53, and 0.17 mg L -1
concentrations at the 120-cm depth measurements found levels >15 mg L
for years 2-4, respectively; only 4 of 29
-1 and 12 of 29 indicated concentrations of 5-10 mg
L -1 . Nitrate concentrations in groundwater, however, were quite low; only 1 of 103
13

NBMA DRAFT Summary - Nitrogen samples indicted a NO
3
- level >10 mg L -1 , and all other were <5.7 mg L -1 . Aschmann et al. (1992) recommended a maximum application rate of ~200 kg N ha -1 to a mature, mixed hardwood forest in Maryland due to the discovery of NO
3 exceeding 10 mg L -1 following applications of >400 kg N ha -1 .
levels in percolate far
The biosolids or residual treatment also may have an affect upon the potential NO
3
leaching rate. At essentially the same solids loading, applications of 1600 kg N ha -1 of fresh biosolids theoretically result in 90 kg N ha -1 NO
3
leaching, while the same amount of aged biosolids should have none (Cole and Henry, 1983). This reflects the effect of N transformations and losses during storage. Medalie et al. (1994) applied biosolids at rates of 199, 396, and 740 kg N ha -1 as total Kjeldahl N (TKN) in 1989 to trenched (no vegetation) and untrenched (with vegetation) plots. After two months, the NO
3
level in soil solution exceeded 10 mg L -1 at all loading rates in trenched plots but was at or below detection limits in untrenched plots for all but the high rate in September, when the NO
3
level was still far below the maximum allowed by the U.S. EPA. The authors concluded that while plant uptake was likely the most important factor in limiting NO
3
leaching, the type of biosolids treatment was also of importance and further suggested that an anaerobically digested, limed, liquid biosolids applied to vegetated lands at rates of up to
14.5 Mg ha -1 is perhaps less likely to lead to NO
3
leaching.
The type, age, size and amount of forest vegetation also has a significant effect upon the NO
3
leaching potential following residual applications. The effect of tree stand age and understory was demonstrated by Cole and Henry (1983). Rates of NO
3
leaching increased as follows: 10 year-old stand with vigorous understory < clear-cut with vigorous understory < older stand with little understory. Following a similar pattern, applications of 47 dry Mg ha -1 to a clear-cut, a 10 year-old stand and a 55 year-old stand resulted in NO
3
leaching of 115, 73 and 480 kg ha -1 , respectively, during the first year, and 7, 5 and 210 kg ha -1 , respectively, during the second year (Henry, 1985). This research lead to conceptual models for biosolids applications to different aged biosolids and different aged Douglas-fir stands. Breuer et al. (1979) found that areas including tree seedlings and native grasses limited NO
3
leaching better than introduced grasses and suggested that a loading rate of 400 kg N ha -1 in treed or grassed areas for wastewater was safe. Varied ages of pine trees were studied by Wells et al. (1986) after the application of liquid anaerobic biosolids. It was found that NO
3
leaching was greatest under stands less than 1 year old followed by 28, 9 and 3 years.
Burton et al. (1990) reported the following trends of NO
3
in groundwater and soil leachate following biosolids application of 22.3 g
N m -2 to 1.5-ha plots of aspen ( Populus sp.), pine ( Pinus sp.) and northern hardwoods
(various species): aspen > pine > northern hardwoods; these results were not predicted by incubated biosolids-treated soil cores. Insam and Merschak (1997) performed a 29-week incubation study of various organic amendments (including both composted [with sawdust or bark] and uncomposted biosolids+shale) utilizing soil cores from a mixed species forest and a spruce forest in Austria. The control plots had initial NO
3
concentrations of 10-40 mg L -1 in leachate, and this increased during the first 10 weeks of the study. A biosolids+shale treatment resulted in NO
10 weeks with peaks of 160 and 140 mg L -1
3
increases in soil water for the first
for the two forest types, respectively, after which it decreased rapidly, but this was followed by an increase in the leachate. After 29
14

NBMA DRAFT Summary - Nitrogen weeks, soil water and leachate of the mixed forest were similar to the control but still higher in the spruce forest (~80 mg L -1 ). For a biosolids+bark (2:1 ratio) treatment, NO
3
in leachate increased during the initial 15 weeks. The biosolids+sawdust (2:1 ratio) treatment resulted in increased NO
3
in soil water only in the spruce forest soil and only during the third week. Brockway and Urie (1983) determined application rates that would not elevate NO
3
levels above 10 mg L -1 analysis; these were: 9.5 dry Mg ha -1 dry Mg ha -1
in northern Michigan by using regression
of undigested biosolids to a red pine plantation, 16.5
of digested biosolids to a red pine and white pine ( plantation, and 19 dry Mg ha -1
Pinus strobus
of digested biosolids to an aspen sprout stand.
L.)
Agricultural applications of biosolids
Variable results concerning NO
3
leaching rates have been found following organic residuals applications to agricultural lands, and, similar to forest applications, the application rate seems to be of paramount importance. Nitrate concentrations increased significantly at the 120 cm depth when biosolids averaging 6.2 percent solids was applied to a 5 cm thickness (King and Morris, 1972). O'Brien and Mitsch (1980) reported increased leaching rates of 21 kg N ha -1 yr -1 for each Mg of biosolids incorporated into farm soil and 22 kg N ha -1 yr -1 for each Mg of biosolids incorporated into stripmine soils.
Intact soil cores were used to determine NO
3
leaching from biosolids vs. antibiotic waste applications by Sommers et al. (1979). It was found that while leaching was significant in soils treated with antibiotic waste, no leaching was noted with biosolids. High levels of nitrates were reported by Soon et al. (1978) in plots treated with ≥1600 kg N ha -1 . Nitrate losses as high as several hundred kg ha -1 were reported when high rates of poultry manure or pig slurry were applied and incorporated into arable soils (Unwin, 1986). Adriano et al.
(1971) measured NO
3
levels as high as 74 ppm in groundwaters beneath an area of densely concentrated dairy farms with mean estimated manure disposal rates of 1000 kg
N ha -1 yr -1 . Following an 11-year study of multiple biosolids applications, Barbarick et al.
(1996) suggested an application of 6.7 dry Mg ha -1 to dryland wheat fields, because that rate resulted in proportionally less N loss and greater soil accumulation and plant uptake of N than a 26.8 Mg ha -1 rate. Additionally, soil profile NO
3
concentrations suggested that leaching losses below 200 cm was greatest at the 26.8 Mg ha -1 rate, and no significant increase in plant growth was found at the higher rate. Ritter et al. (1992) found no significant changes in NO
3 at rates ranging from 66.3 x 10
levels in groundwater following the application of septage
-4 L to 747 x 10 -5 L ha -1 yr -1 and limed with either 3 or 6 g lime L higher NO total of 540 kg N ha ha -1
-1 to corn and fallow fields. Artiola and Pepper (1992b) reported significantly
3
concentrations in soil-pore water in arid, sandy soils on plots amended with a
-1 from biosolids compared to controls, plots amended with 180 kg N
urea, and plots amended with 180 kg N ha -1 liquid biosolids.
The type of biosolids treatment, application method, and time of application also affects the NO
3
leaching potential from agricultural ecosystems. In a two-year study,
Shepherd (1996) found the following order of levels of total N loss following applications of various organic residuals or an inorganic fertilizer at the end of the first year: calcium nitrate > liquid digested injected biosolids > digested liquid surface-applied biosolids > raw liquid injected biosolids > dewatered cake biosolids. The order of N loss based upon time of application was: September (for spring barley) > September (for winter barely) >
15

NBMA DRAFT Summary - Nitrogen
November > January. Results were generally similar during the second year with some changes in the order, however, there was no significant difference between spring and winter barley and total losses of N were smaller. Overall, averaging all applications, there was no significant increase in N concentration in drainage water.
Nitrate leaching greatly depends on the type of crop, primarily due to the amount of N the plant can utilize, and also due to micro-climate conditions. Furrer and Stauffer (1986) found NO
3
leaching was greater on fallow vs. cropped fields, and grasses exhibited the least amount of leaching. Findings that the presence of grass greatly reduced leaching losses as compared to losses under plots with no vegetation were reported by King and
Morris (1974). Kofoed and Klausen (1986) found that leaching after slurry application was highest under crops having a short growing season. The report also stated that beet
(species not provided) roots had a lower leaching rate than did barley. Farm slurry applications on arable versus grassland showed reduced NO
3
leaching under grass in studies done by Steenvoorden (1985). In a report by Unwin (1986), slurry applications to grassland were noted to cause less leaching than applications to uncropped land. The presence of oats decreased the amount of NO
3
leachate in experiments done by Vogt et al. (1981). Studies by Stewart et al. (1975) suggested that applications of biosolids to corn crops be limited to 1.25 cm of 7.3 percent solids due to increased NO
3
leaching at higher application rates. (It should be understood that recommendations such as this are very biosolids-specific.)
General land application of biosolids
Tindall et al. (1994) and Gaggiani (1991) reported extensive results from studies conducted at a long-term, high-rate biosolids land-application site near Denver, CO, with cumulative biosolids applications ranging from 408-634 Mg ha -1 over 17 years; biosolids were either buried or incorporated into the soils. Thirteen wells in the alluvial aquifer contained water high in NO
3
, including three wells downstream from the application site.
The highest NO
3
levels in groundwater were found to occur during periods of the greatest rainfall. While all application areas contributed to N-contamination of groundwater, the section there biosolids had been buried 3-5 m below the surface was determined to be the major source of N in groundwaters. The authors concluded that leachate from biosolidsapplied areas probably caused increased NO
3
plus NO
2
concentrations in alluvial aquifer groundwater. There was no indication of high levels of NO
3
in groundwaters of a bedrock aquifer. Additionally, surface waters did not appear to be affected by biosolids application. Similarly, NO
3
from injected biosolids (0-30 cm below the surface) were reported to have been the likely source of NO
3
rates greater than the U.S. EPA maximum in an aquifer below a biosolids injection field in Nebraska (Spalding et al., 1993).
Soil properties
The properties of the soil are important factors in determining whether nitrates will form and potentially leach into the groundwater. The rate at which these processes happen vary with soil conditions, temperature, pH, aeration, moisture, organic matter, season of the year and soil depth (Stednick and Wooldridge, 1979).
16

NBMA DRAFT Summary - Nitrogen
Soil texture . The NO
3
content of groundwater is usually observed to be higher in areas where water can easily percolate through the soil (Ahtiainen, 1986). Therefore, gravelly or sandy soils will generally have higher concentrations of nitrates below the surface than clayey or silty soils. Clayey or silty soils also encourage denitrification, reducing NO
3
available for leaching. de Haan (1986) reported mean NO in drainage water of 90 mg L -1
Kelling et al. (1977) showed high NO
3
-
3
-
for sandy soil, while clay soil was 24 mg L
concentrations
-1 . Studies by
concentrations following applications of liquid biosolids after 10 weeks in sandy soils, while it took one year to reach the same concentrations in silty loam soils. Mills and Zwarich (1982) reported downward percolation of nitrates to be very slow in fine textured soils, but high NO
3
concentrations were found near the soil surface for several years. Comparisons between impermeable gley and moderately drained loam were made in Ireland by Sherwood (1986). Losses were low in the gley compared to the loam. Nitrates did not reach the 15 cm depth in the gley. Henry (1989) reported little difference in NO
3
concentrations at the 50 cm depth between clayey and sandy sites to which pulp and paper sludges had been applied.
Soil pH . Bugbee (1996) reported higher levels of NO
3
and NO
3
+NH
4
+ leaching in media with lower pH and higher percentages of a yardwaste/biosolids compost. The pH of the media was generally of greater significance than the percentage of compost amendment with respect to leaching.
Seasonal influences
Studies by Furrer and Stauffer (1986) found high NO
3
leaching in rainy months after application. Water input beyond field capacity from October to May caused considerable leaching in studies done by King and Morris (1972). This relationship was also noted by
Melkas et al. (1985) as two seasonal peaks of NO
3
flushing in clay soil. A magnification of NO
3
leaching was reported by Unwin (1986) due to extremely wet conditions following very dry months during which slurry applications were made. Shepherd (1996) found that N losses based upon time of application was greater following September applications of biosolids or inorganic fertilizers compared to November applications, and
November applications led to greater N losses than January applications.
The land-application of biosolids and other organic residuals clearly provides numerous benefits but may also cause environmental problems if not managed properly.
When determining whether to apply residuals and at what rates, a thorough understanding of the N cycle is of extreme importance in ensuring a successful organic residuals landapplication program. Application managers must consider the various steps in the N cycle as well as the many other variables that determine the availability and fate of N when establishing their management plans.
Extensive research has documented greatly enhanced plant growth and N uptake following residual applications to forested and agricultural lands, and the majority of the growth response can be attributed to increased levels of plant-available N. However, there remains a significant environmental concern with respect to the potential leaching of
17

NBMA DRAFT Summary - Nitrogen
NO
3
if the application rate is not made at the agronomic rate for the particular system upon which the application occurs. Studies have indicated that all of the following factors influence NO
3
leaching potential: the N application rate, the type of organic residual, residual treatment process, vegetation (age, density, type), the soil (texture, organic matter, CEC, pH, profile depth, moisture level), climatic conditions (rainfall, temperature, humidity), application technique (trenching, incorporation, surface applications) and N cycle processes (mineralization rate, nitrification rate, denitrification, immobilization, volatilization of NH
3
, leaching of NH
4
+ ).
The simplest way to minimize or eliminate NO
3
leaching is to apply at agronomic rates. Knowing the N cycle and the residual product will assist the manager in determining how much N will likely be available, and understanding the ecosystem of the application site will assist the manager in determining the amount of N likely to be utilized by the system. Utilizing this information will result in a highly successful, environmentally beneficial biosolids or other organic residuals application program.
Reference: Henry, C., and R. Harrison. 1998. Environmental Effects of Biosolids
Management. Trace Metals: Potential for Movement and Toxicity from Biosolids
Application, Effects on Wildlife and Domestic Animals from Biosolids Application,
Air Emissions and Ash Resulting from Incineration of Biosolids, Nitrogen Cycle and
Nitrate Leaching from Biosolids Application, Microbial Activity, Survival and
Transport in Soils Amended with Biosolids, The Fate of Trace Synthetic Organics in
Biosolids Applied to Soil, Runoff Water Quality from Biosolids Application, Effects of Organic Residuals on Poplars. Northwest Biosolids Management Association.
18